Biol Invasions DOI 10.1007/s10530-016-1343-7

ORIGINAL PAPER

The invasive ant, Solenopsis invicta, reduces herpetofauna richness and abundance C. R. Allen . H. E. Birge . J. Slater . E. Wiggers

Received: 29 February 2016 / Accepted: 28 November 2016 Ó Springer International Publishing Switzerland 2016

Abstract Amphibians and reptiles are declining globally. One potential cause of this decline includes impacts resulting from co-occurrence with non-native red imported fire ant, Solenopsis invicta. Although a growing body of anecdotal and observational evidence from laboratory experiments supports this hypothesis, there remains a lack of field scale manipulations testing the effect of fire ants on reptile and amphibian communities. We addressed this gap by measuring reptile and amphibian (‘‘herpetofauna’’) community C. R. Allen U.S. Geological Survey, South Carolina Cooperative Fish and Wildlife Research Unit, Clemson University, Clemson, SC 29634, USA H. E. Birge Nebraska Cooperative Fish and Wildlife Research Unit, and School of Natural Resources, University of Nebraska, Lincoln, NE 68583, USA J. Slater South Carolina Cooperative Fish and Wildlife Research Unit, Clemson University, Clemson, SC 29634, USA E. Wiggers Nemours Wildlife Foundation, 239 Stroban Road, Yemassee, SC 29940, USA Present Address: C. R. Allen (&) U.S. Geological Survey, Nebraska Cooperative Fish and Wildlife Research Unit, School of Natural Resources, University of Nebraska-Lincoln, Lincoln, NE 68583-0711, USA e-mail: [email protected]

response to successful fire ant reductions over the course of 2 years following hydramethylnon application to five 100–200 ha plots in southeastern coastal South Carolina. By assessing changes in relative abundance and species richness of herpetofauna in response to fire ant reductions, we were able to assess whether some species were particularly vulnerable to fire ant presence, and whether this sensitivity manifested at the community level. We found that herpetofauna abundance and species richness responded positively to fire ant reductions. Our results document that even moderate populations of red imported fire ants decrease both the abundance and diversity of herpetofauna. Given global herpetofauna population declines and continued spread of fire ants, there is urgency to understand the impacts of fire ants beyond anecdotal and singles species studies. Our results provides the first community level investigation addressing these dynamics, by manipulating fire ant abundance to reveal a response in herpetofauna species abundance and richness. Keywords Amphibians
Reptiles
Community ecology
Red imported fire ant
Exotic species
Global change drivers

Introduction Amphibians and reptiles are declining worldwide due to multiple, interacting drivers of global change

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(Blaustien et al. 1994; Gibbons et al. 2000; Blaustein and Kiesecker 2002; Todd et al. 2010). In addition to habitat loss, commercial exploitation, chemical pollutants, UV radiation, and emerging diseases, there is growing evidence that co-occurrence with the invasive red imported fire ant (Solenopsis invicta) is a driver of this decline (Blaustein and Wake 1990; Wyman 1990; Wake 1991; Blaustien et al. 1994; Allen et al. 1994; Jablonski 1998; Adams 1999; Belden and Blaustein 2002). Introduced through the port of Mobile, Alabama in the early 1930s, the red imported fire ant spread rapidly throughout the southeastern United States (Lofgren et al. 1975) and remains the dominant ant species (Allen et al. 2004). Red imported fire ants densities can reach 123,000,000 individuals per hectare (Lofgren and Vander Meer 1986; Macom and Porter 1996), and they are effective general scavengers and predators, impacting a diverse range of native taxa (Allen et al. 1994, 2004) ranging from small invertebrates (Porter and Savignano 1990) to large vertebrates including reptiles, amphibians, and mammals (Allen et al. 1997a, b). Specifically, there is growing empirical evidence from laboratory investigations and in situ anecdotal evidence documenting the myriad impacts of red imported fire ants on amphibians and reptiles (collectively ‘‘herpetofauna’’) (Allen et al. 2004). This includes observations of fire ants penetrating and envenomating soft shelled eggs (Mount et al. 1981; Chalcraft and Andrews 1999; Diffe et al. 2010; Newman et al. 2014) and consuming hatchlings and newly metamorphosed juveniles from various herpetofauna species (Landers et al. 1980; Freed and Neitman 1988; Moulis 1997; Conners 1998a, b; Allen et al. 2001, 2004; Parris et al. 2002; Marco et al. 2013). There is also evidence that irritation caused by fire ant stings leads to adaptive or altered behavior (Whiting 1994; Langkilde 2009), reduced growth rates, and immediate (e.g., Montgomery 1996) or delayed (e.g., Langkilde and Freidenfelds 2010) mortality. Fire ants also likely have multiple indirect effects on herpetofauna that are more difficult to ascertain from anecdotal evidence or laboratory studies, such as the reduction of an important invertebrate food source in response to fire ant presence (Porter and Savignano 1990; Suarez et al. 1998). Despite the well-documented impacts of fire ants on reptile and amphibian species through multiple mechanisms and at multiple life stages, there are no field level assessments to

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determine whether these impacts scale up to the population or community level. Amphibians and reptiles in much of the southeastern United States share habitat with red imported fire ants and, due to inherent life history and physiological traits, appear especially vulnerable to impacts. Some reptile and amphibian taxa may be more susceptible than other taxa due to differences in habitat preference, oviparous or viviparous strategies, timing and site of reproduction, extent of fossorial behavior, feeding strategies, and energy requirements (Allen et al. 2004). Additionally, common forestry practices of clear-cutting provide attractive habitat for red imported fire ants, potentially intensifying impacts on herpetofauna bound to or near those sites (Zettler et al. 2004; Wetterer et al. 2007; Todd et al. 2008). Some amphibian species, for example, may be especially vulnerable to fire ant attack. Reliant on temporary rainwater pools made increasingly scarce in recently disturbed areas, juveniles must withstand attack and predation from fire ants as they emerge from these pools (Todd et al. 2008). Whether this natural variation in vulnerability to fire ants is reflected at the herpetofauna community level is a hitherto unanswered question. As reptile and amphibian populations continue to decline globally, reducing this uncertainty becomes increasingly important. We address this gap by assessing in situ herpetofauna community response to the effects of experimentally manipulated fire ant reductions.

Materials and methods We sampled herpetofauna on ten paired, 100–200 ha sites located in Beaufort, Colleton, Jasper and Hampton counties in the coastal plain of southeastern South Carolina. Fire ant invasions were first reported in 1970 in Beaufort, Colleton, and Hampton counties in 1970, and 1963 in Jasper County (Callcott and Collins 1996). Eight of these sites were located on private property, and two were on state-owned land managed by the South Carolina Department of Natural Resources. Dominant land cover consisted of relatively open loblolly (Pinus taeda) or longleaf (Pinus palustris) pine overstory with an understory of largely wire-grass (Aristida stricta) with other native grasses, shrubs, and forbs interspersed. We paired sites within landholdings to reduce heterogeneity contributed by

The invasive ant, Solenopsis invicta, reduces herpetofauna richness and abundance

differences in land cover, land use, and management effects. We randomly chose and treated one member of each pair of sites in May of 2000 and May 2001 to reduce fire ant populations, using hydramethylnon (AmdroÓ/ProBaitÓ) and maintaining the other as a control. Prescheduled private landowner and state directed management practices, such as prescribed burning and tree thinning, continued as planned on the sites. We applied hydramethylnon through an aerial application of insect bait at a rate of approximately 1.5–1.75 kg/ha in May of 2000 and 2001. An application rate of hydramethylnon at 1.68 kg/ha has been estimated to reduce infestations by up to 97% within 3–4 months of treatment (Collins et al. 1992). Spring applications is also reported be most effective in decreasing fire ant infestations, with population levels remaining B75% of pre-treatment for up 8 months after treatment (Collins et al. 1992). We measured fire ant abundance at each site using bait-stations consisting of a 50 9 9 mm petri dish containing filter paper saturated with a sugar-based ant attractant and peanut oil and placed at 10 m intervals along three transects, with 10 bait stations per transect (Lofgren et al. 1961). We sampled the same transects for consistency (2000 and 2001 for post-treatment effects and 1999 for pretreatment/baseline abundance). Baits were collected 1 h after placement, and fire ant abundance indexed for each transect, with 1–10 fire ants on a bait station = 1, 10–100 ants = 2, and over 100 ants = 3. Indices were summed for each transect. We sampled amphibian and reptile communities in spring 2000 (prior to hydramethylnon application on the treatment sites), fall 2000, spring 2001, and fall 2001. During the sampling period, we sampled on a daily basis, employing multiple capture methods in order to effectively capture the diverse taxa encompassed by the ‘‘herpetofauna’’ label. To do this, we used drift fence/pitfall bucket arrays with accompanying funnel traps, coverboard arrays, and a pattern of upright polyvinyl chloride (PVC) pipes. We employed pitfall arrays as described by Gibbons and Bennet (1974), using 50 cm high aluminum flashing. We constructed these arrays in a Y-shape of equal angles, consisting of three 7.5 m legs with 20 L plastic buckets buried flush at ground-level in the center, and at the ends of each leg. The bottom edge of the fencing was buried approximately 10 cm below ground-level by using a small walk-behind gasoline trencher. We

placed funnel traps comprised of 1.27 cm square mesh hardware cloth on either side of the fence midway between the center and outside buckets. Funnel traps were a 30 9 30 9 60 cm rectangular trap with a funnel in each end entering the trap above centerline and extending into it about 5 cm. We constructed these funnel traps to capture those taxa that tend to circumvent pitfall buckets or are able to escape pitfalls. All traps were covered or closed when not in use. Although some sampling biases are associated with these arrays (Gibbons and Semlitsch 1981; Dodd 1991), we assumed biases to be uniform between treatment and control areas. While coverboards are typically 66 9 133 cm sheets of plywood (Grant et al. 1992), for this study we employed 60 9 245 cm sheets of 29 gauge roofing tin due to the use of prescribed fires on these sites. Where necessary, we removed litter and duff so that the boards would be in direct contact with the ground. Cover boards are effective for sampling many snake species, as well as lizards and some toads and frogs. We drove one-meter sections of PVC pipe (diam. = 2 cm) upright into the ground to sample arboreal taxa or those able to escape aforementioned traps by climbing (Moulton et al. 1996). Each of the 10 sites contained four drift fence/ pitfall arrays, 12 coverboards, and 12 PVC pipes. We chose the location of these structures in each site to achieve maximum site coverage for each area while avoiding bias associated with factors such as distances from possible breeding habitats using topographic maps and aerial photos. Once captured, we identified the species, sex (if possible), and age class of each individual (juvenile or adult). We then assigned the individual a mark and released it at the site of capture. Marking methods depended on the taxonomic group involved: we used visible implanted elastomer (VIE) for salamanders (Anholt and Negovetic 1998), notched the marginal scutes with a small file for turtles (Cagle 1939), and implanted snakes with a passive integrated transponder (PIT) tag (Elbin and Burgur 1994; Jemison et al. 1995). We assigned marks to anurans and lizards by toe clipping (Donnelly et al. 1994). Methods were approved by the Clemson University Animal Research Committee. We used a randomized block design analysis of variance for all analyses using the SAS statistical package ANOVA-GLM (SAS Institute 1996), blocking by site to account for any differences of location,

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management, habitat, and pretreatment abundance of herpetofauna and fire ant effects. Fire ant index data were treated individually for each sampling period. Herpetofauna data were analyzed with a repeated measures design to account for the multiple posttreatment sampling periods. Herpetofauna analyses were conducted for all captures combined regardless of species, for reptiles only, for amphibians only, by age class, and for individual species with sufficient captures. We used age class as a variable to assess whether or not reproductive success was compromised by fire ant presence. Shannon-Weiner diversity indices were calculated for each site. We omitted recaptures from all analyses. Population estimates were not possible due to very low samples of most species and insufficient recapture rates. As a result, we based our analyses on relative abundance. Potential differences in detectability are unlikely to significantly affect our results because we blocked by habitat/management and randomly applied treatments; however, it is possible, but unlikely, that fire ant reductions affected detectability on treated versus control sites. Where normality assumptions were violated, we transformed our data. Because of the relatively low statistical power associated with a largescale experiment with 5 replicates, we set P \ 0.10 to be significant so that we were more likely to detect potential biological significance in our experiment.

Results Prior to hydramethylnon treatment, there were no differences in the fire ant abundance indices we measured among the future control and treatment sites (Fig. 1; F1,4 = 0.05, P = 0.83). Following hydramethylnon treatments in May 2000 and May 2001, red imported fire ant population indices on treated sites were significantly reduced relative to the untreated, control sites (Fig. 1). Following the May 2000 hydramethylnon treatment, the treatment sites averaged an index of 1.0, while control sites averaged 15.5 (F1,4 = 53.02, P = 0.002). Six months later in fall 2000, indices averaged 11.2 and 4.3 for control and treatment sites, respectively (F1,4 = 26.84, P = 0.007). After the May 2001 treatment, the treatment sites averaged an index of 2.1, and the control averaged 7.9 (F1,4 = 12.46, P = 0.024). By fall 2001 mean indices were 2.6 in the treatment and

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10.1 in the control plots (F1,4 = 12.70, P = 0.024), indicating consistently reduced fire ants in the treatment sites throughout the duration of our study. Total capture and capture per unit effort data for amphibians and reptiles were not normally distributed (P [ 0.05), so we normalized the data with a squareroot transformation prior to analyses. Prior to hydramethylnon treatment, reptile and amphibian community data showed no significant differences between control and future treatment sites for overall captures (F1,4 = 1.25, P = 0.33), captures per unit effort (F1,4 = 1.26, P = 0.33), or species richness (F1,4 = 0.42, P = 0.55). Following treatment, a total of 1583 reptiles and amphibians representing thirtythree species were captured across all sites (treatment and control), with 299 recaptures, for a total of 1284 individuals marked and released (Table 1). Following treatment, we captured significantly more individuals on treatment versus control plots (F1,4 = 6.4, P = 0.07). We caught an average number of 57.1 individuals on treatment sites versus an average of 28.5 individuals on control sites. To account for differences in trapping effort among sites, capture numbers were converted to a capture per unit effort, in this case number of animals per array-1 day-1 (Fig. 2). As with overall captures, we measured significantly higher capture rates on treated versus control areas, with 2.02 and 1.06 captures per unit effort, respectively (F1,4 = 7.65, P = 0.051). We found higher mean species richness in the treated (7.73) versus control plots (5.0) (F1,4 = 21.41, P = 0.010) (Fig. 2), yet we found no significant difference in Shannon-Weiner diversity indices between treatment (2.007) and control (1.426) (F1,4 = 3.48, P = 0.14). Reptiles We detected more reptile species on treatment areas (12.1 mean captures) versus control (5.9 mean captures) following hydramethylnon treatment (F1,4 = 5.81, P = 0.074). This relationship held after accounting for capture per unit effort (0.41 animals array-1 day-1 on treatment areas and 0.24 animals array-1 day-1 on control areas; F1,4 = 5.77, P = 0.074). Reptile species richness on treated sites responded significantly and positively to fire ant reductions (Fig. 3) (F1,4 = 31.14, P = 0.005), with an average number of 4.07 unique reptile species

The invasive ant, Solenopsis invicta, reduces herpetofauna richness and abundance 20

Control Sites

Index of Average Fire Ant Abubndance

18

Treatment Sites 16 14 12 10 8 6 4 2 0

Fall 1999

Spring 2000

Spring 2001

Time Fig. 1 The effect of aerial application of hydramethylnon (treatment sites) on fire ant abundance prior to (Fall 1999) and post treatment (Spring 2000 and Spring 2001). Abundances here are represented by average indices, with 1–10 fire ants on a bait station = index of 1, 10–100 ants = index of 2, and over 100

ants = index of 3. Ten bait stations were located at 10 m intervals along three transects per site. Indices were summed for each transect and then averaged for treatment and control. Error bars represent ±SE

captured per treated site versus an average 2.47 unique species per control site.

difference in the two age groups’ response to treatment in terms of total captures (F1,4 = 2.32, P = 0.20) and captures per unit effort (F1,4 = 2.32, P = 0.20).

Amphibians There were no significant differences in captures on treatment (45 mean captures) versus control (22.6 mean captures) areas for amphibians post-treatment (F1,4 = 4.49, P = 0.102). Amphibian captures per unit effort, however, were significantly higher on treatment (1.62 animals array-1 day-1) versus control areas (0.83 animals array-1 day-1) (F1,4 = 4.91, P = 0.091). Species richness for amphibians (Fig. 3) was significantly higher on treatment areas than on control areas (F1,4 = 5.12, P = 0.087), with an average of 3.67 unique species per site on treatment areas, and 2.53 unique species per site on control areas. Age class We combined juvenile and subadult classes in order to compare them to adult captures following hydramethylnon treatment. We found no significant

Discussion We documented significant increases in herpetofauna (reptiles and amphibians) abundance and richness in response to reduction of red imported fire ants through hydramethylnon application. This is the first community level investigation into the impacts of red imported fire ants on reptile and amphibians, and our findings corroborate previous anecdotal and single species observational evidence that fire ant impact herpetofauna abundance and richness. Through a field scale manipulation, our results confirm that fire ant impacts incurred upon herpetofauna individuals and populations are detectable at the community level. Although many reptile and amphibian species consume invertebrates vulnerable to hydramethylnon, fire ants are abundant and efficient foragers and outcompete most native insects. We also found evidence that

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C. R. Allen et al. Table 1 Species and individual herpetofauna captured (recaptures omitted) over the duration of the study to determine the effects of fire ant reduction (treated) in the coastal plain of South Carolina Species

Control

Treated

Ambystoma opacum

0

1

Plethodon glutinosus

0

8

Order Caudata

Order Anura Acris gryllus Anaxyrus terrestris Gastrophryne carolinensis Hyla cinerea

5

5

31 177

223 295

52

29

Hyla femoralis

0

2

Hyla squirella

4

13

Lithobates clamitans

24

27

Lithobates sphenocephalus

39

65

Pseudacris brimleyi

0

1

Pseudacris nigrita

0

2

Scaphiopus holbrookii

7

7

24

Order Squamata Suborder Lacertilia Anolis carolinensis

10

Aspidoscelis sexlineata

10

9

Ophisaurus attenuatus

0

1

Plestiodon fasciatus

14

18

Plestiodon laticeps Sceloporus undulatus

1 12

4 15

Scincella lateralis

12

65

Agkistrodon contortrix

1

2

Carphophis amoenus

0

2

Cemophora coccinea

0

1

Coluber constrictor

9

16

Crotalus horridus

0

1

Elaphe guttata

2

3

Elaphe obsoleta

0

1

Heterodon platirhinos

2

3

Storeria occipitomaculata

1

0

Tantilla coronata

9

11

Thamnophis sirtalis

0

2

3

1

Suborder Serpentes

Order Testudines Kinosternon subrubrum Terrapene carolina Totals

123

0

2

425

859

fire ants disproportionately affected some herpetofauna species (although are samples were overly small for strong inductive inference). For example, we found that some amphibian species were more abundant on treated sites, including the northern slimy salamander (Plethodon glutinosus), southern toad (Anaxyrus terrestris), eastern narrow-mouthed toad (Gastrophryne carolinensis), and southern leopard frog (Lithobates sphenocephalus). Others were more evenly distributed among control and treatment sites, such as the American green tree frogs (Hyla cinerea). Physiological differences among species’ response to fire ant stings, different attractiveness to fire ants among species or taxa, variation in natural history and habitat use, or behavior may affect exposure and vulnerability. Some species may be more likely than others to develop adaptive behaviors that counteract the morbidity associated with fire ant exposure. One such example is the apparently learned behavior of fence lizards to ‘‘twitch’’ off stinging fire ants (Langkilde 2009). In the southeastern United States, including our study sites located in South Carolina, fire ant populations are predominantly monogynous (single queen). Monogynous colonies have relatively low mound density (* 50/ha in our study) and spread slowly relative to polygynous (multiple queen) ant populations. In other fire ant infested areas of the United States, especially in the south central region where most investigations of fire ant impacts on wildlife have occurred, polygynous (multiple queen) colonies predominate (Porter 1993). Allen (1993) employed a similar experimental design similar to our own in Texas Coastal prairies in systems dominated by the more densely populated polygynous colony types. Post hydramethylnon treatment, fire ant populations declined by *93% but rebounded with a rapidity that necessitated additional treatment 6 months later. Within 8 months, treated ant populations again rebounded to control levels. Even with this inconstant reduction in fire ants, herpetofauna abundance was approximately double on sites treated with hydramethylnon—although this increase was not statistically significant, likely due to replication constraints in such a large study. While Allen’s (1993) data are somewhat inconclusive, they hint at a response similar to that observed in our study.

The invasive ant, Solenopsis invicta, reduces herpetofauna richness and abundance

a 12 Control Treatment

Average Richness

10

8

6

4

2

0 Fall 2000

Spring 2001

Fall 2001

Time

Average Number of Individuals Caught Array -1 Day -1

b

6 Control Treatment

5

4

3

2

1

0 Fall 2000

Spring 2001

Fall 2001

Time Fig. 2 The effect of red imported fire ant reductions (treatment sites) on average herpetofauna (amphibians and reptiles) species richness (a) and captures per array per day (b). Error bars represent ±SE

The mechanisms surrounding fire ant impacts on herpetofauna communities remain largely uncertain. Our work begins to fill this critical gap, applying a

field manipulation to reveal the positive effect of fire ant reductions on herpetofauna abundance and species richness. While population-level impacts of fire ants

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a Average Number of Individuals Caught Array-1 Day-1

Fig. 3 The effect of red imported fire ant reductions (treatment sites) on the average number of amphibian (a) and reptile (b) captured. Error bars represent ±SE

7

Amphibians 6

Control Treatment

5

4

3

2

1

0 Fall 2000

Spring 2001

Fall 2001

Time

Average Number of Individuals Caught Array -1 Day -1

b

7 Control

Reptiles

Treatment

6

5

4

3

2

1

0 Fall 2000

Spring 2001

Fall 2001

Time

on vertebrates may be chronic and incremental (Allen et al. 1997a, b), we documented a relatively rapid response to fire ant population reductions, suggesting herpetofauna populations have the capability to

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rebound quickly when fire ant impacts are relieved, and that the negative effect size is relatively large. However, given the myriad and likely interactive drivers of global change affecting vulnerable reptile

The invasive ant, Solenopsis invicta, reduces herpetofauna richness and abundance

and amphibian taxa, fire ants are one of many variables with which they must contend. For many herpetofauna, there is a lack of baseline data for population levels, distribution, and habitat requirements. Without these data for comparison, chronic effects of fire ant impacts, such as habitat use change, alteration of dispersal patterns, and overall population declines may go unnoticed until critical thresholds are passed. This high uncertainty also makes it difficult to extrapolate our results and predict the effect of fire ants on herpetofauna abundance and diversity at other locations in the United States or globally, emphasizing the need for widespread, increased herpetofauna monitoring in relation to fire ant abundance. In order to fill this gap it is essential to conduct manipulative field experiments on the community-level effects of fire ants on amphibians and reptiles. Herpetofauna have essential roles in ecosystems including nutrient cycling, disease mitigation, food provisioning for higher trophic levels, genetic diversity, and soil engineering (Hocking and Babbitt 2014). Their contribution to social and ecological systems may be fully realized only when populations are lost. Disentangling the myriad effects of global change on amphibians and reptiles is therefore essential to stemming the global loss of these vulnerable communities and the ecological processes they underpin. Acknowledgements The South Carolina Cooperative Fish and Wildlife Research Unit is jointly supported by a cooperative agreement among the U.S. Geological Survey, the South Carolina Department of Natural Resources, Clemson University, and the Wildlife Management Institute. The Nebraska Cooperative Fish and Wildlife Research Unit is jointly supported by a cooperative agreement between the United States Geological Survey, the Nebraska Game and Parks Commission, the University of Nebraska—Lincoln, the United States Fish and Wildlife Service, and the Wildlife Management Institute. An earlier version of this manuscript was improved by comments from D. Ferraro and W. Mills. Comments from Dr. Kevin G. Smith and three anonymous reviewers greatly improved the manuscript. Any use of trade, firm, or product names is for descriptive purposes only and does not imply endorsement by the U.S. Government.

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