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Impacts of Neonics in New York Water Their Use and Threats to the State's Aquatic Ecosystems

Impacts of Neonics in New York Water
Their Use and Threats to the State’s Aquatic Ecosystems
Pierre Mineau; Pierre Mineau Consulting
Pierre Mineau, PhD
Principal Senior Scientist at Pierre Mineau Consulting
Adjunct Professor, Department of Biology, Carleton University
Subject Editor, Earth Systems and Environmental Sciences, Elsevier Publishing
Contact:; +1 (613) 853-2013
Professional website:
Google scholar link:
Access to open source publications:le/Pierre_Mineau/contributions?ev=prf_act
Foreword ......................................................................................................................................................................... 1
Executive Summary ......................................................................................................................................................... 1
Acknowledgments ........................................................................................................................................................... 2
1. General Information .................................................................................................................................................... 2
1.1. Introduction ....................................................................................................................................................... 2
1.2. Growing Evidence of Neonic Surface and Ground Water Contamination in North America .................................. 2
1.3. What is the ‘No Harm’ Level of Neonics in Water? ................................................................................................ 3
1.3.1. Environmental benchmarks: How does USEPA determine these levels
and how does their approach compare to other jurisdictions? ................................................................................... 3
1.3.2. Choosing between acute and chronic benchmarks. .......................................................................................... 4
1.3.3. Choosing an appropriate benchmark for neonics ............................................................................................ 4
1.3.4. How do we account for neonic mixtures? – the need for a proper cumulative assessment ................................. 5
1.3.5. Drinking water standards ............................................................................................................................... 5
2. New York State and Neonics .......................................................................................................................................... 6
2.1. How Much Are Neonics Used in New York? .......................................................................................................... 6
2.2. What Is the Evidence of Contamination? ........................................................................................................... 10
2.2.1. Surface water samples. .................................................................................................................................. 10
2.2.2. Groundwater and drinking water samples ..................................................................................................... 12
2.3. Are Reported Neonic Levels in New York Water Causing Harm to the Environment or Human Health? ............... 13
3. Conclusions ............................................................................................................................................................... 15
References ..................................................................................................................................................................... 16
This report represents the abbreviated version of a more complete analysis and discussion of neonicotinoid water
contamination in New York and across the country. The reader is referred to the complete report for more details.
Neonicotinoid insecticide (neonic) use in New York has increased markedly in the last 15 years, as it has nationwide. Neonics
now frequently appear in New York surface waters at levels expected to cause significant harm to the state’s aquatic ecosystems.
Substantial reductions in outdoor neonic use are recommended to avoid these harms. More intensive monitoring of both
surface and groundwater is also needed, especially for those compounds not currently monitored, including acetamiprid,
clothianidin, and thiamethoxam.
It is estimated that somewhere between 70 to 76 U.S. tons (141 to 152 thousand pounds) of neonic active ingredients were
used in New York in 2014, but this does not account for consumer use products. It is believed that a high proportion of this
total (86 to 93 thousand pounds) comes from agricultural uses. While imidacloprid was the most heavily used neonic overall
(~76 thousand pounds per year), clothianidin appeared to be the dominant neonic used in agriculture (~38 thousand pounds
per year), even though no outdoor use clothianidin products are approved in the state. This anomaly is explained by the fact
that many crop seeds are coated with clothianidin before planting, and that New York State’s Department of Environmental
Conservation does not currently regulate these treated seeds as “pesticides” under state law.
Residues of the neonic imidacloprid have appeared frequently in New York surface water sampling since 2001. The vast
majority of detections (from 90% to 100% each year) exceed the U.S. Environmental Protection Agency’s (USEPA) and our
chosen 0.01 µg/L benchmark for harm to aquatic ecosystems. Many detections exceed that benchmark by a 10-fold margin.
The highest water concentrations (typically 6-8 µg/L) tend to be from sampling locations where imidacloprid has been
detected over multiple years, suggesting that more frequent monitoring would uncover other significant spikes. Groundwater
detections of imidacloprid have been less frequent, but show maximum values in the same range, suggesting that wetlands
recharged by ground water will be impacted also. On Long Island, the latest (2016-2017) U.S. Geological Survey groundwater
testing reveals detections in ~31% of samples taken—although none above the current USEPA drinking water guideline.
Water sampling for other neonics has been sparse, but given the high runoff potential for clothianidin and thiamethoxam
(both greater than imidacloprid), and their significant use in New York, regular monitoring for these chemicals is needed.
Research on the use of these compounds elsewhere suggests that they are also present in New York surface waters at
biologically damaging levels. More importantly, the aggregate amount of neonics is the relevant concern; research elsewhere
has shown that several neonics often contaminate the same water course.
Scientific research has demonstrated that neonic use results in the loss of invertebrate life in both terrestrial and aquatic
systems with effects that can lead to ecosystem-wide perturbations, diminishing consumer species such as birds, fish,
mammals, and other vertebrates. Detected levels of imidacloprid alone in New York streams exceed levels at which deleterious
effects on stream ecology were observed in other research. The probability that imidacloprid and other neonics are causing
ecosystem-wide damage in New York is very high. Substantial reductions in outdoor neonic use are needed to mitigate further
USEPA drinking water benchmarks remain high for all neonics except thiacloprid, and were not exceeded by the water testing
results reviewed in this report. However, recent research on the toxicity of neonic active ingredients, their breakdown products
or degradates, and the chlorinated by-products they form when passing through normal drinking water treatment processes
creates cause for concern. As new human health research on neonics emerges—particularly on chronic exposures—New York
should monitor water sources for the presence of all neonics (not only imidacloprid) down to detection levels of 0.01 µg/L or
lower, especially in areas where surface waters are used for drinking water and on Long Island.
The author would like to thank Jennifer Sass and Daniel Raichel from the Natural Resources Defense Council (NRDC) for
their contributions and input into this report and to NRDC for providing the financial support to produce it. The author
would also like to thank Robert Warfield, Mike Helms, and Dan Wixted from Cornell University for commenting on section
2.1 of the report. The views expressed in this report are those of the author and do not reflect those of NRDC or any other
1.1. Introduction
Neonicotinoids or “neonics” are a relatively new, but widely used, class of insecticides, approved extensively for home, garden,
and indoor use as well as for agriculture (Jeschke et al. 2010). They are frequently applied as a coating on seeds, meaning that
treatment is applied whether or not it is warranted by pest pressure. As “systemic” compounds, neonics have the ability to
penetrate plant tissues, making the plant—including its sap, nectar, and pollen—toxic to insects. Neonics are highly soluble in
water, which, combined with their persistence in soils, popularity, and prophylactic use, have led to widespread contamination
of the environment (Goulson 2013, Morrissey et al. 2015).
While much has been written about the link between neonic use and the collapse of honeybee and other pollinator
populations around the world, their extensive presence in and potential impacts on aquatic environments have only recently
become better known. This report addresses these issues in the aquatic environment of New York State.
1.2. Growing Evidence of Neonic Surface and Ground Water
Contamination in North America
As the ability to detect neonics in water has grown, so has their frequency of detection. At this point in time, imidacloprid,
the first registered neonic, is still the only neonic routinely looked for in most water monitoring programs; not surprisingly, it
is often found in surface waters. On a national level, the U.S. Environmental Protection Agency (USEPA) recently completed
a review of available U.S. Geological Survey (USGS) and California surface water monitoring data (USEPA 2017a). They
report both high detection frequencies for imidacloprid as well as exceedances of USEPA benchmark levels for ecological
harm (described below) by over a thousand-fold (details in section 2.3.). In New York State, imidacloprid detection in routine
surface water monitoring samples has reached as high as 50% of yearly samples (see section 2.2.).
Although water testing data for the other four major neonic chemicals are
less common, a number of research studies and data reviews have found
that clothianidin and thiamethoxam, where used, are found in surface
water samples also (Mineau and Palmer 2013; Main et al. 2014; Anderson
et al. 2015; Morrissey et al. 2015; Miles et al. 2017, 2018; Struger et al. 2017;
Bradford et al. 2018; Hladik et al. 2015, 2018). These results are hardly
surprising given that several of the newer neonics have a higher potential
to run off to surface waters than imidacloprid—as determined by their
water solubility, ability to bind to soil, and persistence in soils. Pesticide
industry scientists (Chen et al. 2002) introduced a validated indicator of
runoff potential called the ‘Surface Water Mobility Index’ or SWMI, ranging
from 0 (low mobility) to 1 (high mobility). The SWMI values for all the
major neonic chemicals are tabulated in Table 1, along with the pesticide
atrazine—a ubiquitous contaminant of agricultural watersheds—for
Table 1: Surface Water Mobility Indices
(SWMIs) for neonicotinoid insecticides
and atrazine
SWMI Valuea
Atrazine 0.76
Acetamiprid 0.35
Clothianidin 0.66
Dinotefuran 0.85
Imidacloprid 0.56
Thiacloprid 0.30
Thiamethoxam 0.82
a Physicochemical properties obtained from Pesticide Properties
database (PPDB 2018).
As can be seen, three neonics (clothianidin, thiamethoxam and dinotefuran) are expected to be more likely to run off to
surface water than imidacloprid, an already frequent surface water contaminant. Not surprisingly, neonics have been found to
be ubiquitous in tap water drawn from surface waters as well as shallow alluvial wells in areas of heavy neonic use (Klarich et
al. 2017; Sultana et al. 2018).
Ground water studies are fewer and smaller in scope, but evidence of contamination is clear (Mineau and Palmer 2013). As
reviewed by this author, over a decade ago, USEPA (2008) reported imidacloprid levels in New York groundwater as high as
those found in surface streams (7 µg/L; see more complete analysis below). At about the same time, thiamethoxam was also
recorded at even higher levels in Wisconsin wells (Huseth and Groves 2013).
Where detected in water, neonics tend to be present for extended periods of time. For example, in a study of California flowing
water sources, concentrations of imidacloprid were found to remain steady for periods exceeding three months at all sites
monitored throughout the summer (Starner and Goh 2012).
1.3. What is the ‘No Harm’ Level of Neonics in Water?
1.3.1. Environmental benchmarks: How does USEPA determine these levels and how does their
approach compare to other jurisdictions?
To evaluate whether a pesticide may harm aquatic ecosystems, scientists estimate the concentration of that pesticide in water
at or above that at which they would expect ecological impacts to occur. These ‘benchmark values’ try to provide a single
concentration that is sufficiently protective of diverse ecosystems. However, different experts and governments disagree on the
best way to derive these benchmarks.
USEPA has long favored the ‘most sensitive’ species method—setting the benchmark at the lowest acute or chronic toxicity
value from the available species data. Unfortunately, this method often misleads, because USEPA has data on a handful of
species and even closely related species differ greatly in their sensitivity to pesticides. Truly finding the ‘most sensitive’ species
in an ecosystem may require testing hundreds (if not thousands) of different species, which USEPA is never in a position to
do. Indeed, data sets typically available to USEPA include testing results for one to three species only.
Reliance on these small data sets causes benchmark values to fluctuate wildly when new species data become available, as the
history of USEPAs aquatic benchmark for imidacloprid shows. From 1994 to 2017, USEPAs benchmark for acute harm from
imidacloprid became 90-fold more protective and the chronic benchmark 50-fold more protective—taking nearly 25 years to
align these values with those produced by other international efforts. This history, detailed in the full report, also demonstrates
how the benchmark setting processes for “newer” neonics, such as clothianidin, thiamethoxam, and others remain at a very
early stage of development and are likely to follow the same course.
USEPA’s benchmark setting process for estuarine and marine waters—dating back to 2004—also is misguided. Even for
imidacloprid, the most-researched neonic, USEPA admits to a higher uncertainty for its 16.5 µg/L acute and 0.16 µg/L chronic
estuarine/marine benchmarks given the paucity of toxicity data for salt water species (USEPA 2017a). This reliance on a few
arbitrarily-chosen brackish or saltwater species to characterize the entire marine environment is scientifically unsound. A
more credible scientific position is to combine both freshwater and saltwater species and adopt the same benchmark values
for all aquatic environments. This approach has been shown to be scientifically justified (Maltby et al. 2005). Selecting an
appropriate estuarine/marine benchmark has particular relevance for New York, which has a marine coastline and valued
estuaries such as the Hudson River estuary.
Recognizing the shortcomings of the USEPA method, most other jurisdictions now have adopted alternate strategies—
reviewed in detail in the full report. These rely on curve-fitting of toxicity distributions and/or the application of safety factors.
1.3.2. Choosing between acute and chronic benchmarks
Regulatory agencies typically produce two types of benchmarks: acute and chronic. Acute benchmarks attempt to identify
pesticide concentrations expected to cause harms from short-term exposure (hours to days). Chronic benchmarks identify
levels that may cause harm when organisms are exposed for prolonged periods of time, typically measured in weeks or months
and potentially encompassing the full life cycle of exposed species.
For neonics, acute benchmarks do not have much predictive value. Widely reviewed and available research (e.g., Morrissey
et al. 2015) suggests that neonic toxicity is essentially cumulative—meaning the length of exposure is directly related to the
toxicity (Sanchez-Bayo 2009, Tennekes 2010a). Further, several studies have also shown that neonics’ high runoff potential
leads to prolonged watershed exposures, ranging from weeks to months to possibly years. For example, Whiting et al. (2014,
2015) studied the consequences of a clothianidin seed treatment in corn at 0.25 mg/kernel, or 1/5 of the federally-allowed
amount, and found residues in runoff water right until the end of the sampling period—156 days after planting. Similarly,
Schaafsma et al. (2015) found pre-planting levels of both clothianidin and thiamethoxam in ditches and puddles situated in
areas outside seeded fields in central Canada, indicating neonic contamination from the previous growing season (max. of
16.2 µg/L for clothianidin; 16.5 µg/L for thiamethoxam). Cumulative impacts and likely long exposure periods led Pisa et al.
(2017) to conclude in the second iteration of a worldwide, multi-year assessment of neonics and other systemic insecticides
that acute benchmarks are essentially “irrelevant for risk assessments of these chemicals…”
Chronic benchmarks are much more meaningful, although lab testing used by USEPA in benchmark setting may
underestimate actual risks in the field. For example, USEPA (2017a) in their benchmark setting for imidacloprid did not
track whether degradation of the chemical had occurred during the test, even though imidacloprid breakdown compounds
are less toxic to aquatic organisms than the parent molecule (USEPA 2017a). Neonics’ primary breakdown happens through
photolysis—i.e., exposure to light—which is much more likely to occur in laboratory tests, generally performed in clear
water with strong illumination, rather than in a turbid or shaded waterbody more typical in nature. Accordingly, as USEPA
identified, its benchmark setting process may underestimate actual risk in the real world.
1.3.3. Choosing an appropriate benchmark for neonics
As mentioned above, and reviewed in detail in the full version of this report, USEPA benchmark values typically underestimate
risks due to insufficient data and simplifying assumptions. Benchmarks can and do shift dramatically as new data become
available. For example, USEPAs initial acute freshwater benchmark for imidacloprid dropped from 34.5 µg/L in 2007 to
0.385 µg/L in 2017; the chronic benchmark dropped from 0.5 µg/L to 0.01 µg/L over the same period. Recent evaluations by
Canada’s Pesticide Management Regulatory Agency (PMRA) proposed new benchmarks for three major neonics (Table 2).
While there is now some agreement between the two countries’ benchmarks for imidacloprid, there is sizeable discrepancy for
the newer neonics.
Table 2: A comparison of current USEPA aquatic benchmarks to proposed benchmarks by Canada’s PMRA (in µg/L)
Compound USEPA acute
PMRA acute
USEPA chronic
PMRA chronic
Imidacloprid 0.385 0.36b0.01 0.041b
Thiamethoxam 17.5 9.0c0.74 0.026c
Clothianidin 11 1.5d0.05 0.0015d
a USEPA 2019.
b PMRA 2016.
c PMRA 2018a.
d PMRA 2018b.
In an earlier assessment, Mineau and Palmer (2013) proposed that the toxicity of clothianidin and thiamethoxam should
be assumed to be similar to that of imidacloprid because of similar test results when the same species were exposed to the
different neonics. Morrissey et al. (2015) reached similar conclusions, arguing that “differences in relative toxicity among the
various individual neonicotinoids are minor. Indeed, the main difference between imidacloprid and the other newer neonics
is the amount of available toxicity data. Where toxicity data for the same species tested with multiple neonics are available,
they show remarkably similar toxicity.
From this perspective, USEPAs benchmarks for clothianidin and thiamethoxam are untenable. Other benchmarks have
been proposed that better account for the similarities in toxicity among neonics. Morrissey et al. (2015) proposed 0.035 µg/L
as a chronic exposure benchmark for all neonics in order to “avoid lasting effects on aquatic invertebrate communities.
PMRA’s analysis suggests that a scientifically-defensible chronic benchmark for thiamethoxam is slightly lower than that for
imidacloprid, while the clothianidin benchmark should be radically lower at 0.0015 µg/L (Table 2).
In light of these uncertainties, we propose that benchmark values for all neonics should be in the range of 0.01–0.04 µg/L.
For the purpose of this report, we will use the more protective value of 0.01 µg/L given that this is now the value accepted by
USEPA for imidacloprid, although PMRA’s much lower benchmark for clothianidin (0.0015 µg/L) suggests that our chosen
benchmark may not be protective enough.
1.3.4. How do we account for neonic mixtures? – the need for a proper cumulative assessment
Although neonics have similar modes of action, USEPA does not appear to assess their impacts cumulatively, stating that it
has not made “a common mechanism of toxicity” finding for either imidacloprid or thiamethoxam (USDA Forest Service
2016, USEPA 2011). While highly questionable for imidacloprid, this finding is clearly unsupportable when applied to
thiamethoxam, given that the neonic clothianidin is actually the primary breakdown product of thiamethoxam.
Contamination studies in agricultural environments have shown that several neonicotinoid insecticides are often found in
the same water samples. Hladik and Kolpin (2015) in the first USGS national survey of neonics recorded that 26% of their
samples contained more than one neonic and 11% contained three or more.
In an ecological context, there is no valid reason not to consider the registered neonics as a group. Morrissey et al. (2015)
proposed that their benchmarks be applied to the sum of all neonic residues. Bayer Crop Science (2010), the main neonic
manufacturer, has even argued that neonics can act synergistically—meaning that their combined impact might be greater
than the action of each separately. This raises the possibility that we are under-protecting by merely adding concentrations of
the various neonics found in the same water sample.
For these reasons, we propose that the USEPA benchmark of 0.01 µg/L be the benchmark we apply to all neonics separately or
to the combined concentration of neonics in any one sample. This is the level of contamination to which we will compare all
water samples to in order to better predict the ecological harm of the residue levels detected.
1.3.5. Drinking water standards
As early as 1994, when imidacloprid was first registered, USEPA expressed concerns about its potential to contaminate
drinking water supplies (USEPA 1994a, b). Comparing imidacloprid to aldicarb, another highly soluble pesticide, USEPA
found imidacloprid’s potential to leach into groundwater to be three times higher—a finding of particular relevance to New
York given the long history of aldicarb contamination in Long Island aquifers.
Recent research shows USEPAs initial fears were well-founded. Klarich et al. (2017) showed that neonics can survive standard
water treatments and Sultana et al. (2018) found that neonics are near ubiquitous drinking water contaminants in agricultural
areas. Clearly, the human population is now potentially exposed to neonics in a chronic fashion.
USEPA recently updated its guidance document for setting human health drinking water benchmarks (USEPA 2017b). The
process starts by noting the lowest No Observable Adverse Effect Levels (NOAELs) from various mandated toxicological
studies, and deriving appropriate Reference Doses (RfDs) or population adjusted doses (PADs) through the application
of various safety factors reflecting the type and severity of the effects, and completeness of the database. Although USEPA
determines both acute and chronic drinking water benchmarks, we will concentrate here on the chronic benchmarks for the
reasons discussed above.
Table 3 shows the chronic or lifetime reference dose or adjusted reference dose as well as the water concentration that should
not be exceeded in order to remain below this lifetime dose. Included in the table are the toxicology endpoints (adverse
effects) that served as the departure point for these drinking water benchmarks.
Table 3: A summary of current Human Health Benchmarks for Pesticides (HHBPs) in drinking water
or lifetime
or lifetime
in relevant
chronic study
Effect seen in study from which NOAEL establishedc
Imidacloprid 0.057 360 5.7 NOAEL for two year rat study. Thyroid effects at higher doses.
Clothianidin 0.098 630 9.8
NOAEL for two generation rat study. Effects seen include
decreased body weight gain, delayed sexual maturation,
decreased thymus weight in first generation offspring and
increased stillbirths at higher doses.
Thiamethoxam 0.012 77 1.2
NOAEL for two generation rat study. Effects seen on sperm
counts and testis weight in first generation offspring at higher
Thiacloprid 0.004 0.8b1.2
NOAEL for two year rat study. Liver and thyroid
histopathology, nervous system degeneration, and
carcinogenicity seen at higher doses.
Acetamiprid 0.071 450 7.1 NOAEL for two year rat study. Liver and kidney toxicity, body
weight and body weight gain effects seen at higher doses.
a USEPA 2019b.
b Dierent calculation to account for carcinogenicity of compound. e value of 0.8 µg/L is based on an estimated increased cancer risk of one in a million.
c Complete review mammalian toxicology available in Mineau and Callaghan 2018.
The same industry studies were reviewed by other regulatory bodies, notably in California, Canada, and Europe. Concurrence
amongst jurisdictions provides additional weight to the review of the toxicological studies provided by industry, even if the
methods of deriving water concentration benchmarks vary somewhat. However, it should be noted that some independent
published research has resulted in lower NOAELs in some cases (Mineau and Callaghan 2018), but this research has not yet
been incorporated into formal regulatory assessments.
The choice of NOAELs reported above met with general concurrence with the exception of acetamiprid. The European Union
(2004) identified a slightly lower NOAEL of 6.5 mg/kg/day based on a two generation rat study. However, both USEPA and
PMRA identified the NOAEL as 18 mg/kg/day in the same study. This difference is slight and would not lead to a substantial
difference in the resulting drinking water benchmark.
The California EPA (2018) calculated an imidacloprid acute reference water level for non-nursing infants of 283 µg/L. This is
lower than the 360 µg/L level chronic calculated by USEPA. They recommended using this as the official California reference
level for imidacloprid.
Drinking water standards do not take pesticide degradates into account and it is known that some of these are much more
toxic to mammalian systems than the parent material. Recently, Klarich Wong et al. (2019) identified not just the known
metabolites of imidacloprid (imidacloprid-urea and desnitro-imidacloprid) in drinking water supplies, but also four novel
chlorinated products following water treatment. These compounds have not yet been characterized toxicologically and may
present human health concerns.
2.1. How Much Are Neonics Used in New York?
We calculated the total neonic use in New York by using data from USGS (Thelin and Stone 2013); they used confidential
surveys of pesticide use patterns on different crop types and extrapolated these use rates by the recorded crop acreages for
each U.S. state and county. In cases where the USGS use rate data for a county was lacking, we opted for the USGS method of
interpolating from use rates in nearby counties rather than assuming no use—a much more realistic approach. Importantly,
USGS estimates reflect the agricultural use of the pesticides only—not domestic, landscape, or industrial uses—meaning
imidacloprid use is likely greatly underestimated given its extensive use in residential and commercial settings.
We report on three separate years of estimated use: 2004, 2009, and 2014 (Table 4). The year 2004 was chosen because it falls
at the midpoint of intensive surface water sampling for imidacloprid in New York and corresponds with the highest reported
frequency of detection (see section 2.2.1.). The year 2014 is the last year when USGS incorporated seed treatment uses in its
pesticide estimates. This is critical given that imidacloprid, clothianidin, and thiamethoxam are used, to a large extent, in seed
treatments. For example, Figure 1 shows the sharp drop off in estimated thiamethoxam use for 2015 when USGS stopped
tracking seed treatment uses.
Figure 1: National estimatesa of thiamethoxam use, by year and crop
Finally, 2009 was chosen as the midpoint between those two time points to illustrate any trends.
Table 4: USGS-derived estimates of agricultural neonic use for New York State in 2004, 2009, and 2014
2004 quantity (lbs) 2009 quantity (lbs) 2014 quantity (lbs)
Acetamiprid 218 8,265 6,464
Clothianidin 432 17,317 38,561
Dinotefuran 0 0 88
Imidacloprid 4,855 6,279 26,444
Thiacloprid 154 8,234 7,676
Thiamethoxam 2,156 3,893 12,707
Total neonics 7,818 43,989 91,941
This ‘snapshot’ of pesticide use for New York echoes the trends seen worldwide with respect to the increasing dominance of
neonics in insecticide markets (Jeschke et al. 2010). It also suggests that neonic use in New York increased 12-fold in the 2004-
2014 decade, and that clothianidin has become the dominant agricultural neonic, despite the fact that New York has effectively
prohibited non-seed treatment uses of that chemical in agriculture (Serafini 2007).
a Downloaded February 2019 from:
The New York State Department of Environmental Conservation (NYSDEC) also oversees state pesticide reporting
requirements in collaboration with Cornell University, which collects several types of data (PSUR 2019). Prior to 2013, data
were reported by USEPA product registration number, making tabulation difficult without access to the machine-readable
records. Starting in 2013, however, data were amalgamated by active ingredient.
Two NYSDEC data categories are useful to build a full picture of neonic use in New York State (Robert Warfield, Cornell pers.
comm.). The first is “Sales by Commercial Permit Holders” of pesticides intended for agricultural crop production. This data
should best correspond with USGS estimates for agricultural use, with the exception of neonic-treated seeds, which are not
considered pesticides by NYSDEC (Dan Wixted, Cornell University pers. comm.). Given the fact that a large proportion of the
insecticide applied to seeds escape to the environment, either through runoff or during seeding, and that the contamination
of terrestrial invertebrates is higher following seed treatment use than during foliar application (Mineau and Callaghan 2018),
the notion that seed treatment use is not a pesticide application is clearly one based on convenience rather than sound science.
The second is “Use Data by Commercial Applicators, which tallies applications by commercial applicators required to report
annually. Since farmers are exempt from this reporting requirement, this second category should capture non-agricultural
neonic uses. Importantly, however, this number excludes all uses of neonic consumer products (e.g., lawn care products
available at a hardware store or nursery). The other two types of data collected by NYSDEC—“Sales Data to End Users” and
“Sales Data to Resellers”— are likely encompassed in the first two categories.
For comparison, USGS estimates, as well as the estimates generated from the NYSDEC-collected data for the year 2014, are
presented in Table 5.
Table 5: Selected data from NYSDEC for neonic sale and use in 2014
USGS Estimate of
Agricultural Use (lbs)
DEC -Sales by Commercial
Permit Holders (lbs)
DEC - Use Data by
Commercial Applicators (lbs)
Acetamiprid 6,464 1,193 3,911
Clothianidin 38,561 0 4
Dinotefuran 88 0 0
Imidacloprid 26,444 10,842 49,939
Thiacloprid 7,676 7,233 33
Thiamethoxam 12,707 4,314 1,825
Total 91,941 23,583 55,713
Although the “Sales by Commercial Permit Holders” category best estimates agricultural use, instances where farmers contract
out their spraying to a commercial applicator would appear in the “Use Data by Commercial Applicators” category (Robert
Warfield, Cornell pers. comm.). The thiacloprid estimates (USGS vs. NYSDEC-recorded sales) are in good agreement,
suggesting that few non-farmer commercial applicators apply thiacloprid on agricultural land. The case, however, is different
for acetamiprid, suggesting a substantial contribution by non-farmer commercial applicators.
We can estimate the proportion of imidacloprid, thiamethoxam, and clothianidin applied in the form of seed treatments
based on a graphical extrapolation of the USGS use data plots for those chemicals used nationwide from 2010 to 2014
(when seed treatments were included in the estimates) to 2015 (when seed treatments were excluded) and comparing this
extrapolated value to the reported value for 2015 (see full report for details). This provides a rough estimate of the proportion
of the chemicals used in seed treatments–58%, 75%, and 92% for imidacloprid, thiamethoxam, and clothianidin, respectively.
We will assume these proportions hold for New York State.
These estimates allow us to correct the NYSDEC sales data to account for seed treatment use. For example, assuming that
75% of agricultural thiamethoxam use is not captured in the “Sales by Commercial Permit Holders” category, the corrected
total would be approximately 17,196 lbs or 35% higher than the USGS estimated use. Likewise, the corrected agricultural
use of imidacloprid is approximately 25,794 lbs—essentially the same as the USGS estimate. The strong agreement between
NYSDEC-recorded sales and USGS estimates for thiacloprid and imidacloprid and the fair agreement for thiamethoxam
suggest that the USGS estimates for clothianidin are reasonable also, despite the lack of NYSDEC-recorded sales. It is not
surprising that no sales of clothianidin were recorded given that this chemical is almost exclusively used as a seed treatment
and not otherwise permitted for sale or use for agricultural purposes in the state.
One difficulty with sales data is that a sale in any given year does not necessarily reflect use in that same year. However, over
a number of years of data collection, the supply and demand equations should equilibrate to reflect the turnover of various
products. However, this remains an uncertainty.
We believe that the best overall estimate for neonic use in New York is a combination of the USGS and/or NYSDEC estimates
of sale for agricultural use corrected for seed treatment use. The non-agricultural use is best captured by the NYSDEC use
data by commercial applicators. As mentioned earlier, some of these applicators might be applying the pesticides on contract
to farmers. However, the generally strong agreement between USGS estimates and NYSDEC sales data corrected to include
seed treatments and the fact that imidacloprid represents 90% of the neonics applied by commercial applicators suggest that
the “Use Data by Commercial Applicators” category captures the extensive turf, ornamental, and industrial site neonic use
rather than a major part of the agricultural use.1 The recent work by Nowell et al. (2017; but samples were taken in 2013) in
the Midwest confirms that imidacloprid is now being detected more frequently and at higher levels in surface waters with
urban inputs. Based on our assumptions and correcting for seed treatment applications, the only neonic where application
to agricultural land by commercial applicators appears to be happening to any significant extent is acetamiprid. Not knowing
how much of the commercial applications overlap with USGS estimates for agricultural land leads to a higher uncertainty for
this chemical.
Accordingly, Table 6 provides what we believe to be the best estimate of neonic use for New York in 2014.
Table 6: Best estimates of neonic use for New York in 2014 (see text for data sources and assumptions)
Chemical(s) Best estimate (or range) of
agricultural use (lbs)
Best estimate of non-agricultural
use (lbs) (may include some
application to agricultural land
by commercial applicators, for
acetamiprid especially)
Best estimate of neonic use in
New York 2014 (lbs) excluding
any consumer product use
Acetamiprid 1,193 – 6,464 3,911 5,104 – 10,375
Clothianidin 38,561 438,561 – 38,565
Dinotefuran 0 – 88 00 – 88
Imidacloprid 25,814 – 26,444 49,939 75,753 – 76,383
Thiacloprid 7,233 – 7,676 33 7,266 – 7,710
Thiamethoxam 12,707 – 17,258 1,825 14,533 – 19,083
Total Neonics 85,508 – 93,491 55,712 141,217 – 152,205
As can be seen, despite the fact that clothianidin likely predominates in agriculture, imidacloprid remains the dominant
product sold and used in the state. It should be noted, however, that the estimated 70-76 U.S. tons of neonic active
ingredient used in New York excludes consumer products containing neonics. As such, the imidacloprid use totals, especially,
underestimate actual use given the hundreds of imidacloprid consumer products registered for use in the state.
USGS estimates suggest that the use of neonics has increased exponentially in the last decade. As of 2015, NYSDEC reported
that there were 335 imidacloprid products alone registered for use, 252 of which could be used on Long Island (NYSDEC
2015). Based on the same use data source reported in Table 5 above, NYSDEC estimated that the number of applications of
imidacloprid on Long Island had increased from 39,007 in 2003 to 46,316 in 2012. However, based on the combined NYSDEC
data (Sales by Commercial Permit holders; and Use Data by Commercial Applicators—as per Table 5 above) for the period
2013-2016, imidacloprid use may have peaked in 2015 and may now be declining.
As of 2017, the main thiacloprid product (Calypso) appears to have been suspended. According to the NYSDEC list of
registered products, it was not registered for use on Long Island.
1 Nevertheless, it is believed that applications by commercial applicators on farms is high (Dan Wixted, Cornell University, pers. comm.), although this is a general impression
not specic to any particular product.
2.2. What Is the Evidence of Contamination?
We have estimated that at least 70-76 U.S. tons of neonic active ingredients are applied in New York every year, at least based
on 2014 data. We believe that number is probably increasing still (based on national trends)—even if imidacloprid use has
peaked. The next relevant question is whether we can detect any contamination of either surface or groundwater as a result.
2.2.1. Surface water samples.
All relevant and available New York water quality data from the USGS’s National Water Information System (NWIS 2019)
were downloaded in early February of 2019 following guidance provided in Shoda et al. (2018).
USGS started reporting imidacloprid residues in 2000, but with minimum reporting levels (i.e. detection limits) of 0.1 µg/L
only. In 2001, reporting levels dropped to 0.007 µg/L or just below the USEPA and our benchmark for ecological damage
(0.01 µg/L). Unfortunately, reporting levels then changed over time, making detection more difficult. The minimum reporting
level of 0.007 µg/L increased to 0.02 µg/L partway through 2004 and then to 0.06 µg/L partway through 2006. Accordingly,
results for these years were often labeled as trace levels (t), below formal detection levels, and values were mostly given as
estimates. Sampling effort dropped dramatically from 2008 to 2015 with few detections; this interval is therefore presented
as a single time period in the analysis below. From 2013 on, reporting levels were given as 0.011 µg/L on an interim fashion
and then 0.016 µg/L once the use of new calculation software was in place. The current detection levels are above the USEPA
benchmark—clearly inadequate to fully detect harmful biological effects or allow for summation of residues with other
The variation in reporting levels in the last two decades makes it difficult to attach much importance to the proportion of
samples showing detections. However, we can provide a rough estimate: in the years 2001 to 2007 and in 2016, the proportion
of surface water samples analyzed from New York State that showed detections of imidacloprid varied between 15% and 50%
of samples—despite variable minimum reporting levels as reported above. The proportion of imidacloprid detections peaked
at 50% in 2004. Drawing meaningful conclusions from this high level estimate, however, is difficult because the samples lack
context: whether they originated from an area of pesticide use, or from an agricultural, urban, or mixed use environment. In
addition, samples were taken at various times of the year and from various types of waterbodies, including large rivers where
the dilution factor can be extreme.
To understand the actual ecological consequences, it is more meaningful to look at samples where imidacloprid was in fact
detected, indicating some use in the watershed. There is clear evidence from the literature that, where neonics are used
on crops, they will be detected in nearby bodies of water at a high frequency (Morrissey et al. 2015). Reported values of
imidacloprid are summarized in Table 7. The data represent all positive samples reported for any given year, with some sites
being sampled more than once (i.e., the values from which the tabulation was made are not strictly independent).
A number of observations can be made from these results. With the exception of the 2008-2015 period of reduced
sampling intensity, it can be seen that, when detected, imidacloprid was present at levels above our benchmark level for
ecological damage 90-100% of the time. A very high proportion of samples (up to 60% of samples in at least two years) had
imidacloprid levels that were over 10 times this critical benchmark. This suggests that impacts to aquatic invertebrate fauna in
New York State from imidacloprid alone have been substantial.
As high as those values are (relative to biological impact levels), they cannot be assumed to be true maxima. There are serious
limitations to ascertaining maximum levels of contamination from routine water sampling. Samples are often taken from
large streams, after significant dilution has occurred, whereas impacts to aquatic life are expected where most of the aquatic
productivity is taking place—in small drainage ditches and ponds bordering field areas or in small feeder streams. Grab or
spot samples taken at a single moment in time are also unlikely to capture peak concentrations. Indeed, it has been shown
that, even when taken weekly, water samples will underestimate peak concentrations by one to three orders of magnitude
(Xing et al. 2015). Many of the New York sampling sites we examined for this report were sampled only once in the 15-year
Table 7: A summary of New York State imidacloprid detections from the USGS NWIS database
Ye a r Number of
High value
Proportion of
values above
the 0.01 µg/L
of values at
least 10 fold
higher than
the 0.01 µg/L
2001 50 0.197 0.132 1.30 98% 60%
2002 31 0.238 0.077 1.84 97% 45%
2003 93 0.271 0.074 7.94 99% 42%
2004 83 0.347 0.085 4.93 99% 60%
2005 63 0.373 0.092 4.64 100% 48%
2006 57 0.144 0.029 5.13 100% 12%
2007 49 0.146 0.033 1.40 90% 22%
2008-2015 8 0.025 0.018 0.073 62% 0%
2016 122 0.082 0.041 0.460 93% 26%
Where a water source was sampled more often, higher levels of imidacloprid were most often detected. Figure 2 shows how
the ability to detect higher levels of contamination in surface waters is directly related to the intensity of sampling. Here, the
maximum observed concentration of imidacloprid at any given site (as high as 7.9 µg/L on one site) is plotted against the
number of years imidacloprid was detected at that site.
0 2 4 6
Years detected
8 10
Max concentration
Expon. (Max concentration)
The shape of the curve in Figure 2 suggests that
high values of imidacloprid detected in routine
water sampling are less a function of the specific
site sampled, but rather, reflect sites with more
intensive sampling. If correct, this means that most
sites with imidacloprid detections will be exposed
to these very high and extremely damaging levels of
Outside of imidacloprid, the sampling for other
neonics has been minimal. Based on the same
NWIS data, a total of only fifteen samples were
analyzed for clothianidin and thiamethoxam over
the same 2001-2016 period—all from between 2012
and 2016—with reporting limits between 0.0039 to
0.0062 µg/L for clothianidin and 0.0034 to 0.0039
µg/L for thiamethoxam. Only one of the fifteen
samples was taken in June, and therefore probably
occurred post-seeding; this sample from Fall Creek
near Ithaca contained 0.020 µg/L clothianidin and
0.0119 µg/L thiamethoxam, as well as trace levels of
Figure 2: Maximum concentration of imidacloprid (in µg/L)
detected at a sampling site against number of years of detection
at that site with exponential fit shown
The last three neonics, acetamiprid, dinotefuran, and thiacloprid were analyzed only at three New York sites. Most of the
sampling consisted of a monthly sample taken from the Genesee River in Rochester in 2015-2016. Minimum reporting levels
were low (0.0032-0.0045 µg/L), but there were no detections from this limited sampling effort.
The low monitoring effort for neonics other than imidacloprid is unfortunate in light of the longer persistence and greater
potential of the newer products to contaminate surface waters. Hladik et al. (2015) found that both clothianidin and
thiamethoxam occurred more frequently than imidacloprid in Midwest streams fed from agricultural areas.
Apart from the regular USGS testing, the U.S. Fish and Wildlife Service (USFWS) conducted a standalone water testing project
in New York (Secord and Patnode 2018). Eleven surface water samples were taken in July 2016 from streams downstream
from, or in proximity to, potato fields with a limit of detection of 0.004 µg/L. For this sampling, four of eight sampled
locations tested positive for neonics. Table 8 provides a summary of those detections.
Table 8: A summary of positive neonic detections in Secord and Patnode (2018)
Sample ID Analytes Concentration (µg/L) Site description
Salmon Creek 1, Wayne
County, NY - Downstream
of potatoes, corn, beans,
NYP19 Imidacloprid
Salmon Creek 2, Wayne
County, NY - Downstream
of corn, beans, potatoes,
NYP20 Imidacloprid 0.006
Flint Creek, Yates County -
Downstream of potatoes,
corn, beans
NYP23 Imidacloprid 0.018
Cohocton River, Steuben
County, NY - Downstream
of potatoes, corn, beans,
a Average of two measurements.
Of these, three of the four sample locations had neonics above the 0.01 µg/L benchmark, with the combined neonic
concentration at one site exceeding the benchmark by a factor of 30. Half of the sites with positive detections had residues
of multiple neonics.
In the same project, USFWS also took five aquatic invertebrate samples described as “mostly crayfish.” Given the high
detection limit of 3 µg/L, only one of five invertebrate samples tested positive for neonics: 7.3 µg/L of thiacloprid. This is
an interesting finding in light of the low estimated amount of thiacloprid used in New York.
Finally, Carpenter and Helbling (2018) included imidacloprid as one of a large number of “micro-pollutants” analyzed in the
Hudson River Estuary. Of the 127 water samples taken in 2016-2017 from 17 sites, imidacloprid was detected in 36 samples
with a median concentration of 0.012 µg/L and a maximum of 0.103 ug/L; therefore, slightly over half of the detected levels
were above the biological harm benchmark. Interestingly, these authors found that imidacloprid detections tended to cluster
with samples affected by wastewater outlets (and sewage treatment plants in particular), again suggesting the importance of
urban landscape uses in the overall contamination picture for that neonic.
2.2.2. Groundwater and drinking water samples
Based on the same search of NWIS records, 1,023 groundwater samples were tested by the USGS for imidacloprid between
2001 and 2016, with detection limits as described above for surface water samples. Of those, only 46 detected neonics, ranging
from trace levels to 5.3 µg/L. All but three detections had values above the ecological harm benchmark, but none approached
the drinking water benchmark set by USEPA.
An exchange between NYSDEC (NYSDEC 2003, 2004) and Bayer Corporation (Bayer Corp. 2004) mentions the detection of
imidacloprid in approximately 20 monitoring and private wells on Long Island. Imidacloprid was reported from well clusters
down gradient from farms and, in some cases, trees injected with imidacloprid, with the highest reported value at 0.0067 µg/L.
There were concerns expressed over the fact that some wells with detections were well away from use sites and others were
deep (85-90 feet) community supply wells.
In 2014, NYSDEC (NYSDEC 2014) summarized drinking water quality on Long Island based on analyses carried out by the
Suffolk County Department of Health Services and Water Authority between the years 1996-2010 (Table 9). Imidacloprid was
the only neonic analyzed. It was the sixth most frequently detected pesticide in Long Island groundwater, whether as part of
the public water system or in private wells.
Table 9: Imidacloprid data (µg/L) from the Suffolk County Department of Health Services (1996-2010)
from NYSDEC (2014)
Drinking water supply Monitoring wells
of detects
Minimum Maximum Median Number of
detects Minimum Maximum Median
549 (60) 0.1 12.9 0.4 341 0.04 407 0.7
Imidacloprid urea, a degradate, was also detected in six monitoring well samples at a maximum of 1.3 µg/L. The guanidine
and olefin degradates were not detected.
A 2015 NYSDEC (NYSDEC 2015) report summarized ground and well water data from Long Island from 2001 to 2013. This
report cited 0.2 µg/L as the lowest concentration detected (at odds with the previous report). Using that higher detection
limit, the proportion of wells where imidacloprid was detected in the last year of sampling (2013) was 8.4% for monitoring
wells and 7.5% for private wells with maximum levels of 6.2 µg/L and 2.8 µg/L respectively. That same report pointed out that
detection levels from 2010 to 2013 were lower than from 2005 to 2009.
The latest USGS information available for Long Island (2016-2017) provides results for 88 groundwater samples. Twenty-
seven of those samples (31%) had detections of imidacloprid ranging from 0.005 µg/L to a maximum of 5.3 µg/L.
NYSDEC with the help of Cornell University investigated well water for pesticide contamination in several upstate New York
counties from 2006 to 2010 (Whitbeck 2006, 2008, 2009a,b, 2010). Wells were chosen on the basis of intensity of nearby
pesticide use and vulnerability. Imidacloprid was analyzed from year two of the program onwards, but with a detection limit
of 1 µg/L only. However, in one of the years, samples were tested with a commercial enzyme-linked immunosorbent assay
(ELISA) for imidacloprid and related compounds (degradates as well as thiacloprid and acetamiprid) with a detection limit of
0.07 µg/L and quantification limit of 0.2 µg/L. This resulted in a single detection estimated to be between those two values.
2.3. Are Reported Neonic Levels in New York Water Causing Harm to the
Environment or Human Health?
The data available for New York indicate that, when detected, neonic levels are likely to be above the ecological harm level of
0.01 µg/L, and frequently above 0.1 µg/L. There already is an indication that, with more intensive sampling, contamination
levels above 1 µg/L would regularly be encountered. Levels of 8 µg/L have already been seen in routine monitoring efforts,
which, as discussed above, are unlikely to detect true maximum pesticide concentration levels. Directed studies (reviewed in
the full-length version of this report) have shown much higher concentrations of neonics measured in and around treated
The USEPA (2017a) has already concluded that, nationwide, imidacloprid levels are frequently above the level (0.01 µg/L) at
which they believe some aquatic invertebrate species will be harmed. In fact, they conclude that several key taxonomic groups
of aquatic invertebrates, not merely the most sensitive ones, are likely to be adversely affected (USEPA 2017a). They arrived
at this conclusion after reviewing nationwide imidacloprid contamination values as well as modeled water concentrations.
The latter approach allowed them to differentiate between different types of use for imidacloprid. USEPA (2017a) thus
estimated that 60% of seed treatment applications, 90% of soil applications, and 100% of foliar applications of imidacloprid
are expected to produce surface water contamination levels above the 0.01 µg/L benchmark for ecological harm. Moreover, the
loss of dust from treated seeds at planting, an important route of environmental contamination (see Mineau and Callaghan
for an extensive review), was not included in the estimated water contamination levels following seed-treatment use. The
60% estimate for the proportion of seed-treatment uses that produce contamination levels above the aquatic benchmark is
therefore a clear underestimate.
Morrissey et al. (2015) following an exhaustive review of water measurements worldwide similarly concluded that their higher
estimated injury level of 0.035 µg/L was exceeded of the neonic, not just imidacloprid.
Nowell et al. (2017) were able to show the consequences of imidacloprid detection on mayfly abundance in their monitored
streams. Mayflies are known to be sensitive to neonics and are a key component of the benchmark levels established in all
jurisdictions. Although a correlation is not proof of causation, their data suggests that neonics are indeed having deleterious
effects on stream ecology in the Midwest. In that study, the highest recorded concentration of imidacloprid was 2.2 µg/L.
Because concentrations as high as 8 µg/L have been recorded in New York streams, the presumption of ecological damage in
New York is very high. Given the levels at which imidacloprid was detected in groundwater, impacts to streams and wetlands
receiving groundwater recharge are also likely.
Figure 3: Relationship between mayfly (ephemeroptera) abundance and maximum imidacloprid concentrations in
Midwest streams according to Nowell et al. 2017 (in ng/L or 1/1,000th of 1 µg/L)
Hallman et al. (2014) demonstrated possible broader ecological impacts of neonic contamination by showing a convincing
correlation between neonic use and declining insectivorous bird populations in the Netherlands. They divided the time period
of their analysis into pre-and-post-neonic periods, showing that not only did neonic concentrations explain bird declines,
but these site-specific declines were not seen before the introduction of neonics, despite the use of other insecticides of high
aquatic toxicity. Regional bird declines were observed starting at estimated neonic concentrations of about 0.2 µg/L.
Scientists and independent experts are now convinced of the broad impact of the whole class of neonics on ecosystems.
Curiously, the USEPA (2008) in one of its early reviews of thiamethoxam predicted “structural and functional changes of both
the aquatic and terrestrial ecosystems,” an unprecedented assessment and prophetic warning. Others followed suit, including:
Tennekes (2010b), a Dutch scientist and naturalist who predicted a “disaster in the making;” the “Task Force on Systemic
Pesticides, a large group of independent scientists under the auspices of the International Union for the Conservation of
Nature (IUCN) (van der Sluijs et al. 2015, Pisa et al. 2015, 2017); the European Academies Science Advisory Council (2015)
made up of 29 independent scientists nominated by their respective countries; as well as several other independent scientists
(Morissey et al. 2015; Sánchez-Bayo et al. 2016).
Regulatory agencies including Canada’s PMRA, and the European Food Safety Authority (EFSA) agree that imidacloprid as
currently allowed to be used, presents unacceptable risks to the aquatic environment. USEPA (2017a) concludes that their
“risk findings … are in general agreement” with those two agencies. Clothianidin, thiamethoxam, and imidacloprid were re-
evaluated recently by PMRA, which recommended a complete phase out of all outdoor uses of these three neonic on food and
feed crops, including seed treatments and outdoor ornamentals, due to the evidence of serious harm to aquatic species and
ecosystems (PMRA 2016b; PMRA 2018a, b). Likewise, the European Union banned outdoor uses of the same three neonics at
the end of 2018—although this was largely as a result of pollinator impacts. USEPAs re-evaluation of these neonics had not
concluded at the time of publishing this report.
Clearly, New York State is not unique. A broad scientific consensus from around the world suggests that the levels of neonic
contamination reported in New York are indicative of clear damage to aquatic environments and terrestrial consumers, such
as birds, amphibians, and mammals.
Neonic use in New York has increased dramatically in the last two decades, and neonics now frequently appear in the state’s
water supplies. Given the very high probability that the neonic levels found are causing significant harm to aquatic ecosystems,
substantial reductions in outdoor neonic use are recommended.
Imidacloprid residues appear frequently in New York surface water testing, with nearly all detections exceeding USEPA’s
and our chosen 0.01 µg/L benchmark for harm to aquatic ecosystems, and many exceeding it by a 10-fold margin or more.
The highest water concentrations (typically 6-8 µg/L) tend to be from sampling locations where imidacloprid has been
detected over multiple years, suggesting that more frequent monitoring would uncover other significant spikes. Imidacloprid
groundwater detections, while less frequent, were common on Long Island—with 2016-2017 USGS testing revealing the
chemical in ~31% of samples.
Sparse sampling for the other five neonic chemicals provides limited insight into their presence in and effect on New York
water, but given their significant use and high runoff potential, regular monitoring for these chemicals is recommended.
While the aggregate neonic presence in water is the relevant measure of concern, imidacloprid data alone indicate a very high
probability that neonics are causing ecosystem-wide damage—including depletion of aquatic invertebrate populations as well
as possible harms to consumer species, such as birds, fish, and mammals.
Although detected imidacloprid levels did not exceed current USEPA drinking water standards, emerging research on neonics
and their toxic breakdown products in drinking water, as well as the lack of comprehensive monitoring for the other neonics
suggests that risk to the human population has been underestimated. New York should monitor water sources for the presence
of all neonics (down to detection levels of 0.01 µg/L or lower), especially where surface waters are used for drinking water and
on Long Island or other areas with heavy neonic use and vulnerable aquifers.
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... Since the underground water resources are shielded from photolysis, the residues accumulate in the groundwater threatening human life (Op de Beeck et al. 2017). Only limited studies are available on groundwater contamination by neonicotinoids (Mineau and Palmer 2013;Mineau 2019;Blanchoud et al. 2019). Hence, research incorporating pesticide fate models are essential to determine the degradation process and infiltration rate of neonicotinoids, followed by their persistence in the groundwater to divulge the scientific gaps. ...
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The prophylactic use of neonicotinoids in paddy fields has raised concern due to its toxicity to ecological systems and human health. The present study evaluated the concentrations of neonicotinoids such as clothianidin, imidacloprid, thiamethoxam, acetamiprid, and thiacloprid in the water-soil systems of the paddy fields, and their potential discharge into the groundwater along the Cauvery delta region, South India. Though neonicotinoids are extensively sprayed in the paddy fields, the concentration of residues analyzed by QuEChERS, combined with LC–MS/MS found no detectable residues at concentrations above LOD. The LOD and the LOQ values for water and soil were 0.001 ppm and 0.0025 ppm and 0.025 ppm and 0.05 ppm respectively. The results of the study found that neonicotinoids are less persistent in the water-soil systems of the delta region as they are readily exposed to photolysis and undergo rapid microbial degradation. Further, the hydropedological characteristics of the highly saturated delta soil facilitate ready leaching followed by vertical migration and infiltration into the soil aquifers.
Technical Report
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A risk assessment of bats and neonicotinoid insecticides.
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Neonicotinoids are a popular and widely-used class of insecticides whose heavy usage rates and purported negative impacts on bees and other beneficial insects has led to questions about their mobility and accumulation in the environment. Neonicotinoid compounds are currently registered for over 140 different crop uses in the United States, with commercial growers continuing to rely heavily on neonicotinoid insecticides for the control of key insect pests through a combination of in-ground and foliar applications. In 2008, the Wisconsin Department of Agriculture, Trade and Consumer Protection (DATCP) began testing for neonicotinoids in groundwater test wells in the state, reporting detections of one or more neonicotinoids in dozens of shallow groundwater test wells. In 2011, similar detection levels were confirmed in several high-capacity overhead center-pivot irrigation systems in central Wisconsin. The current study was initiated to investigate the spatial extent and magnitude of neonicotinoid contamination in groundwater in and around areas of irrigated commercial agriculture in central Wisconsin. From 2013–2015 a total of 317 samples were collected from 91 unique high-capacity irrigation wells and tested for the presence of thiamethoxam (TMX), a neonicotinoid, using enzyme-linked immunosorbent assays. 67% of all samples were positive for TMX at a concentration above the analytical limit of quantification (0.05 μg/L) and 78% of all wells tested positive at least once. Mean detection was 0.28 μg/L, with a maximum detection of 1.67 μg/L. Five wells had at least one detection exceeding 1.00 μg/L. Furthermore, an analysis of the spatial structure of these well detects suggests that contamination profiles vary across the landscape, with differences in mean detection levels observed from landscape (25 km), to farm (5 km), to individual well (500 m) scales. We also provide an update of DATCP’s neonicotinoid monitoring in Wisconsin’s shallow groundwater test wells and private potable wells for the years 2011–2017.
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New information on the lethal and sublethal effects of neonicotinoids and fipronil on organisms is presented in this review, complementing the previous Worldwide Integrated Assessment (WIA) in 2015. The high toxicity of these systemic insecticides to invertebrates has been confirmed and expanded to include more species and compounds. Most of the recent research has focused on bees and the sublethal and ecological impacts these insecticides have on pollinators. Toxic effects on other invertebrate taxa also covered predatory and parasitoid natural enemies and aquatic arthropods. Little new information has been gathered on soil organisms. The impact on marine and coastal ecosystems is still largely uncharted. The chronic lethality of neonicotinoids to insects and crustaceans, and the strengthened evidence that these chemicals also impair the immune system and reproduction, highlights the dangers of this particular insecticidal class (neonicotinoids and fipronil), with the potential to greatly decrease populations of arthropods in both terrestrial and aquatic environments. Sublethal effects on fish, reptiles, frogs, birds, and mammals are also reported, showing a better understanding of the mechanisms of toxicity of these insecticides in vertebrates and their deleterious impacts on growth, reproduction, and neurobehaviour of most of the species tested. This review concludes with a summary of impacts on the ecosystem services and functioning, particularly on pollination, soil biota, and aquatic invertebrate communities, thus reinforcing the previous WIA conclusions (van der Sluijs et al. 2015).
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Neonicotinoid insecticides are widespread in surface waters across the agriculturally intensive Midwestern United States. We report for the first time the presence of three neonicotinoids in finished drinking water and demonstrate their general persistence during conventional water treatment. Periodic tap water grab samples were collected at the University of Iowa over 7 weeks in 2016 (May–July) after maize/soy planting. Clothianidin, imidacloprid, and thiamethoxam were ubiquitously detected in finished water samples at concentrations ranging from 0.24 to 57.3 ng/L. Samples collected along the University of Iowa treatment train indicate no apparent removal of clothianidin or imidacloprid, with modest thiamethoxam removal (∼50%). In contrast, the concentrations of all neonicotinoids were substantially lower in the Iowa City treatment facility finished water using granular activated carbon (GAC) filtration. Batch experiments investigated potential losses. Thiamethoxam losses are due to base-catalyzed hydrolysis under high-pH conditions during lime softening. GAC rapidly and nearly completely removed all three neonicotinoids. Clothianidin is susceptible to reaction with free chlorine and may undergo at least partial transformation during chlorination. Our work provides new insights into the persistence of neonicotinoids and their potential for transformation during water treatment and distribution, while also identifying GAC as a potentially effective management tool for decreasing neonicotinoid concentrations in finished drinking water.
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The widespread usage of neonicotinoid insecticides has sparked concern over their effects on non-target organisms. While research has largely focused on terrestrial systems, the low soil binding and high water solubility of neonicotinoids, paired with their extensive use on the landscape, puts aquatic environments at high risk for contamination via runoff events. We assessed the potential threat of these compounds to wetland communities using a combination of field surveys and experimental exposures including concentrations that are representative of what invertebrates experience in the field. In laboratory toxicity experiments, LC50 values ranged from 0.002 ppm to 1.2 ppm for aquatic invertebrates exposed to clothianidin. However, freshwater snails and amphibian larvae showed high tolerance to the chemical with no mortality observed at the highest dissolvable concentration of the insecticide. We also observed behavioral effects of clothianidin. Water bugs, Belostoma flumineum, displayed a dose-dependent reduction in feeding rate following exposure to clothianidin. Similarly, crayfish, Orconectes propinquus, exhibited reduced responsiveness to stimulus with increasing clothianidin concentration. Using a semi-natural mesocosm experiment, we manipulated clothianidin concentration (0.6, 5, and 352 ppb) and the presence of predatory invertebrates to explore community-level effects. We observed high invertebrate predator mortality with increases in clothianidin concentration. With increased predator mortality, prey survival increased by 50% at the highest clothianidin concentration. Thus, clothianidin contamination can result in a top-down trophic cascade in a community dominated by invertebrate predators. In our Indiana field study, we detected clothianidin (max = 176 ppb), imidacloprid (max = 141 ppb), and acetamiprid (max = 7 ppb) in soil samples. In water samples, we detected clothianidin (max = 0.67 ppb), imidacloprid (max = 0.18 ppb), and thiamethoxam (max = 2,568 ppb). Neonicotinoids were detected in >56% of soil samples and >90% of the water samples, which reflects a growing understanding that neonicotinoids are ubiquitous environmental contaminants. Collectively, our results underscore the need for additional research into the effects of neonicotinoids on aquatic communities and ecosystems.
The objective of this study was to identify sources of micropollutants in the Hudson River Estuary (HRE). We collected 127 grab samples at seventeen sites along the HRE over two years and screened for up to 200 micropollutants. We quantified 168 of the micropollutants in at least one of the samples. Atrazine, gabapentin, metolachlor, and sucralose were measured in every sample. We used data-driven unsupervised methods to cluster the micropollutants based on their spatiotemporal occurrence and normalized-concentration patterns. Three major clusters of micropollutants were identified: ubiquitous and mixed-use (core micropollutants); sourced from sewage treatment plant outfalls (STP micropollutants); and derived from diffuse upstream sources (diffuse micropollutants). Each of these clusters was further refined into sub-clusters that were linked to specific sources based on relationships identified through geospatial analysis of watershed features. Evaluation of cumulative loadings of each sub-cluster revealed that the Mohawk River and Rondout Creek are major contributors of most core micropollutants and STP micropollutants and the upper HRE is a major contributor of diffuse micropollutants. These data provide the first comprehensive evaluation of micropollutants in the HRE and define distinct spatiotemporal micropollutant clusters that are linked to sources and conserved across surface water systems around the world.
Because of the persistence and solubility of neonicotinoid insecticides (NNIs), there is concern that these compounds may contaminate sources of drinking water. The objective of this project was to evaluate the distribution of NNIs in raw and treated drinking water from selected municipalities that draw their water from the lower Great Lakes in areas of southern Ontario, Canada where there is high intensity agriculture. Sites were monitored using Polar Organic Chemical Integrative Samplers (POCIS) and by collecting grab samples at six drinking water treatment plants. Thiamethoxam, clothianidin and imidacloprid were detected in both POCIS and grab samples of raw water. The frequency of detection of NNIs was much lower in treated drinking water, but some compounds were still detected at estimated concentrations in the low ng L-1range. Thiamethoxam was detected in one grab sample of raw drinking water at a mean concentration of 0.28 μg L-1, which is above the guidelines for drinking water recommended in some jurisdictions, including the European Union directive on pesticide levels <0.1 μg L-1in water intended for human consumption. Further work is required to determine whether contamination of sources of drinking water with this class of insecticides is a global problem in agricultural regions.
To better characterize the transport of neonicotinoid insecticides to the world's largest freshwater ecosystem, monthly samples (October 2015-September 2016) were collected from 10 major tributaries to the Great Lakes, USA. For the monthly tributary samples, neonicotinoids were detected in every month sampled and five of the six target neonicotinoids were detected. At least one neonicotinoid was detected in 74% of the monthly samples with up to three neonicotinoids detected in an individual sample (10% of all samples). The most frequently detected neonicotinoid was imidacloprid (53%), followed by clothianidin (44%), thiamethoxam (22%), acetamiprid (2%), and dinotefuran (1%). Thiacloprid was not detected in any samples. The maximum concentration for an individual neonicotinoid was 230 ng L-1 and the maximum total neonicotinoids in an individual sample was 400 ng L-1. The median detected individual neonicotinoid concentrations ranged from non-detect to 10 ng L-1. The detections of clothianidin and thiamethoxam significantly increased as the percent of cultivated crops in the basins increased (ρ = 0.73, P = .01; ρ = 0.66, P = .04, respectively). In contrast, imidacloprid detections significantly increased as the percent of the urbanization in the basins increased (ρ = 0.66, P = .03). Neonicotinoid concentrations generally increased in spring through summer coinciding with the planting of neonicotinoid-treated seeds and broadcast applications of neonicotinoids. More spatially intensive samples were collected in an agriculturally dominated basin (8 sites along the Maumee River, Ohio) twice during the spring, 2016 planting season to provide further information on neonicotinoid inputs to the Great Lakes. Three neonicotinoids were ubiquitously detected (clothianidin, imidacloprid, thiamethoxam) in all water samples collected within this basin. Maximum individual neonicotinoid concentrations was 330 ng L-1 and maximum total neonicotinoid concentration was 670 ng L-1; median detected individual neonicotinoid concentrations were 7.0 to 39 ng L-1.
Aquatic organisms in streams are exposed to pesticide mixtures that vary in composition over time in response to changes in flow conditions, pesticide inputs to the stream, and pesticide fate and degradation within the stream. To characterize mixtures of dissolved-phase pesticides and degradates in Midwestern streams, a synoptic study was conducted at 100 streams during May–August 2013. In weekly water samples, 94 pesticides and 89 degradates were detected, with a median of 25 compounds detected per sample and 54 detected per site. In a screening-level assessment using aquatic-life benchmarks and the Pesticide Toxicity Index (PTI), potential effects on fish were unlikely in most streams. For invertebrates, potential chronic toxicity was predicted in 53% of streams, punctuated in 12% of streams by acutely toxic exposures. For aquatic plants, acute but likely reversible effects on biomass were predicted in 75% of streams, with potential longer-term effects on plant communities in 9% of streams. Relatively few pesticides in water—atrazine, acetochlor, metolachlor, imidacloprid, fipronil, organophosphate insecticides, and carbendazim—were predicted to be major contributors to potential toxicity. Agricultural streams had the highest potential for effects on plants, especially in May–June, corresponding to high spring-flush herbicide concentrations. Urban streams had higher detection frequencies and concentrations of insecticides and most fungicides than in agricultural streams, and higher potential for invertebrate toxicity, which peaked during July–August. Toxicity-screening predictions for invertebrates were supported by quantile regressions showing significant associations for the Benthic Invertebrate-PTI and imidacloprid concentrations with invertebrate community metrics for MSQA streams, and by mesocosm toxicity testing with imidacloprid showing effects on invertebrate communities at environmentally relevant concentrations. This study documents the most complex pesticide mixtures yet reported in discrete water samples in the U.S. and, using multiple lines of evidence, predicts that pesticides were potentially toxic to nontarget aquatic life in about half of the sampled streams.