Stream Habitats and Aquatic Communities in an Agricultural Watershed: Responses to a Mandatory Riparian Buffer Law

doi:10.21203/rs.3.rs-2625011/v1

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Stream Habitats and Aquatic Communities in an Agricultural Watershed: Responses to a Mandatory Riparian Buffer Law

https://doi.org/10.21203/rs.3.rs-2625011/v1

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published 28 Jul, 2023

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Riparian buffers along streams can intercept eroding soils, contaminants, and nutrients, improving stream habitats and increasing the health of aquatic communities. Instream and riparian habitats and fish and benthic invertebrate communities were surveyed in a Minnesota stream draining an agricultural watershed before and after implementation of a state-mandated buffer law passed in 2014 and aimed at protecting water quality. Intensive habitat assessments, electrofishing, and benthic invertebrate sampling were used at the same 13 sites in 2005 and 2018. Average buffer width nearly doubled between surveys, and instream abundance of fine sediments and embeddedness of coarse substrates by fine sediments both declined significantly within 1 to 3 years of buffer establishment. Stream sites also were significantly deeper with faster current velocities, and sites had increased riffle habitat and increased instream vegetative cover for fish after buffer mandates. However, fish and invertebrate biotic integrity scores, and other biotic community metrics, did not display significant improvements after buffer establishment. Stream habitats appear to respond quickly when intact and continuous riparian buffers insulate streams from surrounding agricultural activities, but improvements in biotic communities likely will require more time to adapt to changed habitat conditions.

Riparian buffer

agricultural stream

fish

macroinvertebrates

stream habitats

Agricultural landscapes can impact their local rivers and streams via eroding soils (Armour et al. 1991; Waters 1995), stormwater runoff containing agricultural fertilizers and other chemicals (Battaglin and Fairchild 2002; Schäfer et al. 2007, Tesoriero et al. 2009; Dubrovsky and Hamilton 2010), and altered runoff hydrology (Freeman et al. 2007; Blann et al. 2009). These factors in turn may degrade stream habitats (Davis et al. 2003), change instream productivity (McTammany et al. 2007), reduce or eliminate sensitive aquatic taxa (Lenat 1984), and alter aquatic food webs (Moss 2008). Such impacts may be highly localized (Nerbonne and Vondracek 2001) or more widespread (Heitke et al. 2006), at times homogenizing habitats and communities across broad geographical areas (Li and Reynolds 1994).

Local soils and geology can significantly alter the impacts of agriculture on rivers and streams (Reynolds et al. 1986; Troelstrup and Perry 1989; Molenat et al. 2008). For example, flat topography allows precipitation falling on farmlands to be filtered and purified by soils and underlying bedrock before reaching aquifers or local rivers (Fragalà and Parkin 2010; Hénault-Ethier et al. 2017). In contrast, the combination of steep topography, thin hydrophobic soils, and porous (or karst) geology may allow agricultural soils and chemicals to reach aquifers and surface waters quickly and unmodified (Troelstrup and Perry 1989; Boyer and Pasquarell 1996, 1999; Lindsey et al. 2010; Danz et al. 2013; Roissiez et al. 2013).

Vegetated riparian buffers can insulate lotic systems from land use effects (Talmage et al. 2002; Vondracek et al. 2005; Yates et al. 2007). Wide, intact forest and grass buffers can intercept eroded soils in runoff before they reach rivers and streams, ultimately reducing nutrient concentrations in both surface runoff and subsurface waters (Osborne and Kovacic 1993; Barker et al. 2006). Protection from potential stream contaminants can allow diverse and high-quality stream communities to exist even in watersheds dominated by agriculture (Wang et al. 1997; Fitzpatrick et al. 2001; King et al. 2005). In the United States, the use of vegetated riparian buffers along waterways is expanding to better manage chemical runoff and soil erosion from agricultural lands (United States Department of Agriculture 2016). Unfortunately, the often patchwork implementation of buffers along most rivers and streams has reduced their overall effectiveness (Barker et al. 2006).

Stream biotic communities (e.g., fish, benthic invertebrates) in agricultural regions can respond to development of riparian buffers in several ways. Buffers have been shown to improve fish and invertebrate community integrity and diversity (Barker et al. 2006; Duehr et al. 2008), expand food resources for fish (Albertson et al. 2018), and enhance fish growth (Fischer et al. 2010). However, such observations are not universal, and many studies have reported little to no effect of riparian buffers on stream communities flowing through agricultural areas (e.g., Sovell et al. 2000; Weigel et al. 2000; Smiley et al. 2011). Lack of biotic response to buffers has been attributed variously to differing influences of conservation practices at different spatial scales (Weigel et al. 2000; Fischer et al. 2010), presence of agricultural drainage tiles allowing continued delivery of nutrients to the stream (Smiley et al. 2011), lack of sensitive or intolerant fish or invertebrates to recolonize impacted stream reaches (Fischer et al. 2010; Mundahl and Mundahl 2022), insufficient time since buffer establishment (Sovell et al. 2000; Duehr et al. 2008), and a continuing, overwhelming influence of agriculture on the watershed (Smiley et al. 2011).

A recurring problem with assessments of buffer effects on river and stream environments is a distinct lack of before-and-after data on stream habitat and biotic community response to buffer establishment (Duehr et al. 2008). Buffer studies typically have made concurrent comparisons of reaches with and without buffers on the same stream (Duehr et al. 2008; Mundahl and Mundahl 2022), reaches with different buffer types on the same or different streams (Sovell et al. 2000; Weigel et al. 2000), a variety of streams with and without buffers (Barker et al. 2006; Fischer et al. 2010; Smiley et al. 2011; Albertson et al. 2018), or streams with buffers differing in time since establishment (Duehr et al. 2008). Assessments of the same stream reaches before and after buffer establishment are rare (Duehr et al. 2008).

In an effort to protect surface water quality by expanding and improving on the use of streamside buffers on a broad scale, the State of Minnesota passed a law in 2014 (Minnesota Statutes 2014, section 103B.101, subdivision 12, as amended in 2016 and 2020; https://www.revisor.mn.gov/statutes/cite/103F.48 ) that mandated the presence of a continuous, vegetative buffer (perennially rooted vegetation, 15-m average width, 9-m minimum width) between all public waters and cultivated lands. Within vegetated buffers, haying and livestock grazing are permitted activities. To help meet the requirements of that law, the United States Department of Agriculture and the State of Minnesota agreed in 2017 to spend $500 million over 5 years to protect and improve waters in 54 counties primarily by installing > 12,000 hectares of vegetated buffer strips along streams, rivers, and lakes (Minnesota Board of Water and Soil Resources 2017). The final deadline for establishing buffers along public waters was 1 November 2017.

The goal of our study was to examine the influence of a newly established buffer law on instream and riparian habitat, and on aquatic community structure, in a Minnesota, USA, stream flowing through an agricultural landscape. Specifically, we investigated whether the mandated establishment of stream-side buffers reduced the impacts of agricultural activities on stream habitats 1 to 3 years after buffer compliance, and whether buffer establishment resulted in improvements in fish and/or benthic invertebrate community structure. We used data gathered from the same stream sites prior to and after buffer rule implementation to assess possible changes. We hypothesized that we would observe increased buffer widths in tandem with reductions in the presence of fine stream substrates (sand, silt, clay), reduced embeddedness of coarse substrates by fine substrates, reduced bare soil on banks (with increased grass and forb coverage), increased water depths, and reduced stream widths. In turn, we predicted that reduced fine sediments and suspected reductions in nutrient inputs would lead to increased abundances of intolerant fish and benthic invertebrate taxa, reductions in tolerant species, and overall improvements in fish and benthic invertebrate community integrity scores.

Study area

The South Fork Whitewater River (SFWR) lies within the Driftless Area ecoregion in southeastern Minnesota, USA. This area has been an important agricultural area for > 160 years due to fertile, loess soils overlying ancient (> 100,000 ybp) glacial till (Williams and Vondracek 2010). The region lies within the transition zone between eastern hardwood forests and western tallgrass prairies (Whitewater River Watershed Project 2015).

Poor land use (e.g., forest removal, plowing steep slopes, continuous cropping, burning and grazing hillsides) in southeastern Minnesota resulted in severe soil erosion during the 1920s and 1930s (Whitewater River Watershed Project 2015). The Whitewater River watershed was catastrophically impacted by soil erosion and flooding, as mud and logging debris from the watershed clogged streams. This 830-km² drainage suffered through many multi-flood years (e.g., 28 flood events in 1938; Waters 1977), leading to a major federal project during the mid-1940s to construct erosion control structures and perform forest management work within the watershed. These conservation measures and improving land stewardship led to significant improvements within the drainage basin, but degraded water quality continues to plague the present-day Whitewater River system (Whitewater River Watershed Project 2015).

The Whitewater River remains impacted by agricultural activities; most river sections are impaired for aquatic recreation (due to high fecal coliform bacteria) and aquatic life (due to turbidity; U.S. Environmental Protection Agency Clean Water Act, Section 303(d) listing of impaired waters), with elevated (> 10 mg/l) nitrate-nitrogen concentrations common throughout the system (Minnesota Pollution Control Agency 2013). Aquatic biota (fish and macroinvertebrate communities) in the Whitewater River may be limited further by warm water temperatures, low dissolved oxygen, and/or degraded stream habitats (Schauls and Carter 2015).

The SFWR drains the southern and eastern portions (241 km², 29%) of the Whitewater River basin in Olmsted and Winona counties in southeastern Minnesota, USA (Fig. 1), an area dominated by agriculture (49% cropland, 28% livestock pastureland; Minnesota Pollution Control Agency 2013). From its headwaters 387 m above sea level, the SFWR flows 56.5 km to the east and north, descending 168 m through an oversized valley carved by glacial meltwaters to its confluence with the mainstem Whitewater River. The SFWR watershed also includes > 100 livestock feedlots that generate > 600 metric tons of manure/day, applied as fertilizer on watershed croplands at a rate of 15 metric tons/hectare/year (Minnesota Pollution Control Agency 2013). More than 4300 ha (23%) of agricultural lands within the SFWR watershed utilize various conservation practices (e.g., reduced tillage, cover crops, nutrient management, grass waterways) to reduce potential impacts to stream water quality (Minnesota Pollution Control Agency 2022). Despite these efforts, the entire SFWR is listed as impaired for aquatic recreation and aquatic life (Minnesota Pollution Control Agency 2013).

Field Work

During the summers (June to August) of 2005 (prior to state buffer laws) and 2018 (after state buffer laws), we surveyed 13 stream sites along the SFWR (Fig. 1) for fish, aquatic macroinvertebrates, and stream habitat. Sites were spaced irregularly along the stream due to lack of road access in some locations and private property access restrictions in others. Sites were located on both public and private property, encompassing croplands, livestock pastures, parklands, and forests.

At each site, fish were sampled with one or more backpack electrofishing units (Smith-Root LR-24, 12-B POW, and Type VII) as needed, based on stream size. All fish collected within a 150-m stream reach (downstream to upstream, single pass) were identified, counted, and released. Occasional specimens of uncertain identity were retained for later identification. This sampling procedure was the standard protocol recommended when using a coldwater Index of Biotic Integrity (IBI; Mundahl and Simon 1998) to assess fish communities.

Benthic aquatic macroinvertebrates were collected with a D-frame net (500 µm mesh) at each stream site. Each sample was a composite of all organisms collected by disturbing the substrate upstream from the net (0.1-m² area) for 30 seconds in each of two sections (fast and slow) of a single riffle. Three such composite samples were collected at each site, each from a separate riffle. When riffles were absent, collections were made from any available coarse (gravel, cobble, boulder) substrates. Macroinvertebrate samples were preserved in 70% ethanol and later sorted, identified (mostly to genus level), and counted in the laboratory. This sampling protocol followed the procedure recommended when using a regional Benthic Index of Biotic Integrity (B-IBI; Wittman and Mundahl 2003; Magner et al. 2008) to assess benthic aquatic macroinvertebrate communities.

Physical habitat at each stream site was examined in both 2005 and 2018 using detailed habitat assessments based on protocols developed or adapted for use regionally (Platts et al. 1982; Simonson et al. 1993; Minnesota DNR 2007). Sampling at each site was conducted using a modification of the transect method (Simonson et al. 1993). Habitats at each site were assessed across each of 12 to 15 transects spaced 10 m or more apart. Along each transect, instream habitat features were measured, and riparian conditions were recorded. Instream measures (depth, current velocity, dominant substrate, substrate embeddedness) were taken at four, evenly spaced points along each transect. At each point, depth was measured using a wading rod, and velocity measured at 0.6-depth (Marsh-McBirney model 2000 flow meter). Dominant substrate composition was estimated visually according to a modification of the Wentworth Scale (clay, silt, sand, gravel, cobble, boulder; Minnesota DNR 2007). Percent fines were determined by comparing the number of observations of clay, silt, and sand combined relative to all substrate observations. Embeddedness, the percent of large substrates such as cobbles covered by fine materials, was visually estimated and scored on a five-category scale: score of 1 = < 5% embeddedness, 2 = 5–25%, 3 = 26–50%, 4 = 5–75%, and 5 = > 75% (Platts et al. 1983).

Other instream measures were estimated visually within a reach one mean stream width in length, centered on the transect (Simonson et al. 1993). These included percentage of the stream shaded by the riparian canopy at noon, percentages of riffle, pool, and run, and percent of the reach providing vegetative cover for fish 200 mm or larger, all estimated to the nearest 5%.

Riparian measures were made at one stream bank per transect, alternating the side measured with each transect. Width of the riparian buffer was measured to the nearest meter (with a meter tape or laser rangefinder). Average length of vegetation overhanging the stream was measured to the nearest 0.1 meter. The percentages of bank vegetation as grass, forb, tree, and shrub, and the percentage of bank as bare soil or rock, were estimated visually to the nearest 5% for each category. Data collected from all transects were averaged to determine overall site values.

Data Analyses

Prior to comparing stream habitat and biotic data between 2005 and 2018, we assessed fish and aquatic macroinvertebrates in SFWR by calculating community integrity indices for each group (which compared them to regional standards). Fish data were used to calculate a coldwater IBI score and rating for each site (Mundahl and Simon 1998), based on the presence of brown trout throughout most of the SFWR. Twelve fish community metrics (five species richness metrics: total, coldwater specialists, minnows, benthic, tolerant; five percentage metrics: salmonids as brook trout, intolerant individuals, coldwater individuals, white suckers, top carnivores; and two abundance metrics: number of coldwater individuals, number of warmwater individuals) were quantified, scored, and summed to produce a coldwater IBI score for each site (scale = 0 to 120; 105–120 = excellent, 70–100 = good, 35–65 = fair, 10–30 = poor, 0–5 = very poor). These ratings allowed us to compare the fish communities observed at SFWR sites to regional benchmarks (Mundahl and Simon 1998). Using this coldwater IBI, ratings of poor and very poor indicate high to severe impairment, with degraded conditions and trout very rare or absent. In contrast, fair to excellent ratings indicate reduced impairment and trout abundance ranging from common to abundant.

Benthic invertebrate data were used to calculate regional B-IBI scores and ratings for each stream site (Wittman and Mundahl 2003; Magner et al. 2008). The index was developed based on mostly genus- and family-level identifications (Wittman and Mundahl 2003). Ten invertebrate metrics (seven taxa richness metrics: total, Plecoptera, Trichoptera, Diptera, long-lived, intolerant, filterer; three percentage metrics: Plecoptera individuals, predators, long-lived individuals) were quantified, scored, and summed to produce a B-IBI score (scale = 0 to 100; 65–100 = excellent, 50–60 = good, 30–45 = fair, 10–25 = poor, 0–5 = very poor) for each individual sample. Scores for the triplicate invertebrate samples for each site were averaged to produce a single value and rating for that site. These ratings allowed us to compare the benthic invertebrate communities observed at SFWR sites to regional standards (Wittman and Mundahl 2003; Magner et al. 2008). With this B-IBI, ratings of fair to very poor indicate moderate to severe impairment, with degraded conditions producing reduced taxa richness (especially Ephemeroptera, Plecoptera, and Trichoptera) and increasing dominance by tolerant taxa. In contrast, good to excellent ratings indicate slight or no impairment and high taxa richness (including intolerant or sensitive taxa).

We used a combination of visual and statistical methods to examine fish and benthic invertebrate data before and after implementation of the buffer law. First, we conducted principal component analysis (PCA; JMP Pro 16 software) on abundances of the various taxa to visually depict possible shifts in either community from before to after the buffer law. Next, we used a series of simple paired t tests to compare total fish and benthic invertebrate IBI scores, as well as individual metric values, between years for sites sampled in both 2005 and 2018. We used a Bonferroni correction for multiple comparisons using the same data to set alpha = 0.0038 for each test for fish comparisons and 0.0045 for benthic invertebrate comparisons. These comparisons (13 for fish, 11 for invertebrates) allowed us to assess the effects of the buffer law on the entire fish and invertebrate community integrities, as well as on various subcategories of each community. Percentage data were log-transformed prior to testing to meet normality assumptions.

Instream and riparian habitat data were summarized prior to comparing data collected before and after the buffer law. This involved averaging multiple measurements (e.g., 60 depth measurements) for most variables to produce a single mean value for each year’s survey at each site for comparison between years. We then used both a visual PCA analysis and a series of 18 simple paired t tests to compare habitat values between years for sites sampled in both 2005 and 2018, to assess the effects of the buffer law on stream habitat. Percentage data were log-transformed prior to testing to meet normality assumptions. We set alpha = 0.05 for all habitat variable comparisons.

Eleven of the 18 instream and riparian habitat variables examined exhibited significant change after versus before implementation of mandatory buffers within the SFWR watershed (Table 1). As hypothesized, average buffer widths increased significantly at the study sites, nearly doubling in size compared to their pre-law size. In 2005, seven of the 13 sites had minimal to no buffers, none of which met the 15-m width standard mandated in 2014. In 2018, buffer widths at those same seven sites averaged 90 m, six-fold greater than the mandated standard, whereas buffer widths at the remaining sites were unchanged. In 2018, buffers at all 13 sites were in compliance with the state buffer law.

Table 1

Mean (± SD) instream and riparian habitat variables for 13 sites in the South Fork Whitewater River before and after buffer law implementation. Significant differences (P < 0.05) are highlighted with bold italics
Variable	Before	After	t value	P
Stream width (m)	5.7 (3.3)	6.8 (2.9)	3.36	0.003
Water depth (cm)	35 (12)	42 (12)	1.21	0.127
Current velocity (cm/sec)	16 (8)	36 (9)	9.32	< 0.001
Percent fines	53 (26)	43 (17)	2.13	0.030
Embeddedness	4.6 (0.7)	2.1 (0.6)	7.99	< 0.001
Percent shaded	19 (17)	26 (27)	1.59	0.071
Percent riffle	11 (12)	23 (15)	2.57	0.014
Percent run	75 (18)	67 (13)	1.17	0.145
Percent pool	14 (13)	10 (12)	0.67	0.259
Percent vegetative cover	11 (14)	21 (11)	3.03	0.006
Buffer width (m)	55 (73)	100 (60)	2.185	0.027
Overhanging vegetation (cm)	53 (26)	18 (17)	3.78	0.002
Percent bare bank	21 (12)	50 (31)	3.69	0.002
Percent bank grass	59 (21)	43 (39)	4.17	< 0.001
Percent bank forbs	17 (11)	8 (5)	2.85	0.009
Percent bank shrubs	1 (2)	1 (2)	1.09	0.150
Percent bank trees	1 (1)	2 (2)	1.69	0.061
Percent bank rocks	2 (4)	5 (11)	1.32	0.108

After riparian buffers were mandated, several other stream habitat variables at SFWR sites also changed significantly (Table 1). For example, instream abundance of fine sediments and embeddedness of coarse substrates by fine sediments both declined significantly as hypothesized (prevalence of fines decreased by nearly 20%, embeddedness reduced by 50% or more). Stream sites also had significantly faster current velocities, increased riffle habitat, and increased instream vegetative cover for fish after buffer establishment. Contrary to expectations, stream widths were significantly wider after the buffer law, bare soil on banks was proportionally more abundant, and streambank vegetation was reduced compared to pre-buffer law surveys (Table 1).

PCA analysis of stream habitat variables in 2005 and 2018 suggest a consistent change in stream habitats after buffer law implementation at all sites examined (Fig. 2). Stream sites clustered differently between years, separated primarily on PCA axis 1. After the buffer law, sites in 2018 generally had wider buffers, faster currents, deeper water, proportionally less run habitat, and reduced substrate embeddedness and fine sediments compared to the same sites in 2005 before the buffer law.

We collected and identified large numbers of fish and aquatic invertebrates during our SFWR surveys. A total of 8,113 fish from 17 species and 10,163 invertebrates representing 36 taxa were surveyed during the 2005 and 2018 sampling seasons combined. Individual sites averaged nine to 10 species of fish (Table 2) and nine to 10 taxa of benthic invertebrates (Table 3) each year. Common fishes included brown trout (Salmo trutta), slimy sculpin (Cottus cognatus), brook stickleback (Culaea inconstans), creek chub (Semotilus atromaculatus), and white sucker (Catostomus commersonii), central stoneroller (Campostoma anomalum), and western blacknose dace (Rhinichthys obtusus), while amphipods (Gammarus sp.), baetid mayflies (Ephemeroptera, family Baetidae), midge larvae (Diptera, family Chironomidae), hydropsychid caddisflies (Trichoptera, family Hydropsychidae), and riffle beetles (Coleoptera, family Elmidae) were the most common invertebrates.

Table 2

Mean (± SD) fish IBI scores and individual metric values for 13 sites in the South Fork Whitewater River before and after buffer law implementation. No significant (P < 0.0038 due to Bonferroni correction) differences were detected
Variable	Before	After	t value	P
Fish IBI score	13 (18)	17 (20)	1.62	0.066
Number of species	9 (3)	10 (3)	0.19	0.425
Number of coldwater species	1 (1)	2 (1)	1.72	0.055
Number of minnow species	6 (2)	5 (2)	1.88	0.042
Number of benthic species	4 (1)	4 (1)	0.43	0.336
Number of tolerant species	3 (1)	4 (1)	0.82	0.214
Percent salmonids as brook trout	0 (0)	0 (0)	0.00	1.000
Percent intolerant individuals	6 (20)	3 (7)	0.60	0.281
Percent coldwater individuals	19 (32)	16 (20)	0.37	0.357
Percent white suckers	20 (12)	19 (11)	0.26	0.401
Percent top carnivores	5 (16)	10 (13)	1.26	0.115
Number of coldwater individuals per 150 m	53 (103)	34 (49)	1.25	0.118
Number of warmwater individuals per 150 m	331 (192)	206 (189)	1.91	0.040

Table 3

Mean (± SD) benthic invertebrate IBI scores and individual metric values for 13 sites in the South Fork Whitewater River before and after buffer law implementation. No significant (P < 0.0045 due to Bonferroni correction) differences were detected
Variable	Before	After	t value	P
Benthic IBI score	21 (9)	23 (9)	0.21	0.418
Number of taxa	10 (3)	9 (1)	1.05	0.159
Number of Plecoptera taxa	0.1 (0.1)	0.3 (0.4)	2.56	0.014
Number of Trichoptera taxa	2 (1)	2 (1)	1.08	0.152
Number of Diptera taxa	3 (1)	2 (1)	1.23	0.123
Number of long-lived taxa	1 (1)	1 (1)	0.85	0.207
Number of intolerant taxa	1 (1)	1 (1)	0.07	0.473
Number of filterer taxa	2 (1)	2 (1)	1.11	0.146
Percent Plecoptera individuals	0.1 (0.1)	0.6 (0.9)	2.25	0.024
Percent predator individuals	7 (11)	9 (9)	0.72	0.243
Percent long-lived individuals	4 (10)	1 (2)	1.10	0.149

In contrast to significant changes observed in stream habitats, fish and benthic invertebrate community measures at SFWR sites changed only slightly, if at all, after buffer establishment (Tables 2 and 3). Improvements in fish and benthic invertebrate IBI scores were slight, but insignificant (both communities rated as “poor” by their respective IBIs), and none of the community metrics changed after buffer placement. None of the changes expected in tolerant, intolerant, or benthic species were realized. Consequently, neither fish IBI score (r = 0.236, P = 0.268) nor benthic IBI score (r = 0.120, P = 0.575) were correlated with riparian buffer width, and fish and benthic IBI scores were not correlated significantly (r = -0.158, P = 0.462) with one another.

PCA analyses of fish and benthic invertebrate communities before and after buffer law implementation (Fig. 3) corroborated the lack of statistical changes in biotic community metrics. There were no clear differences in either fish or invertebrate community compositions at sites sampled pre- and post-buffer law.

Minnesota’s mandatory buffer law eliminated the previous patchwork distribution of streamside buffers within the state, with most counties in nearly total compliance (99%+) with the law (Minnesota Pollution Control Agency 2022). Our study found that several instream habitat features (e.g., % fine substrates, substrate embeddedness) responded as anticipated to expanded riparian buffers at 13 stream sites on the SFWR, whereas other instream and riparian variables (e.g., water depth, stream width, % bare soil on banks) did not respond, or changed in unexpected ways. In contrast, neither fish nor benthic invertebrate communities responded to the observed changes in instream habitats that accompanied the improved buffers.

Intact vegetative buffers can insulate streams from adjacent agricultural activities by trapping eroded soils, and any chemicals adsorbed to them, that may be carried by the overland flow of storm waters (Osborne and Kovacic 1993; Talmage et al. 2002; Vondracek et al. 2005; Yates et al. 2007). Reduced delivery rates of fine soil particles, and continued flushing of fine particles by elevated streamflows during storm events and spring snowmelt, potentially reduced the instream abundance of sand, slit, and clay, while also reducing the embeddedness of coarse substrates by fine sediments in the SFWR. In addition, the higher current velocities observed in the SFWR post-buffer law should continue to prevent or reduce deposition of fines, as higher and more variable current velocities are likely to produce greater substrate heterogeneity (Duehr et al. 2006). When coupled with the increasing regional stream baseflows in the midwestern United States (Ayers et al. 2018, 2021a, b), greater current velocities could increase the relative abundance of gravel, cobble, and boulder stream-bottom substrates with reduced embeddedness within the SFWR.

Observing positive changes in physical stream habitat only 1 to 3 years after buffer establishment at several SFWR sites was encouraging. The SFWR and other regional rivers and streams were devastated by poor land use practices a century ago, and many of those waters still bear the scars from that abuse (Trimble 2013; Stout et al. 2014). Some streams can recover quickly from agricultural impacts once riparian buffers are in place to insulate streams from their watersheds (Hunter 1991). For example, stream buffers in nearby Wisconsin and Iowa have been correlated with relatively more coarse substrates, fewer fines, and reduced embeddedness (Lyons et al. 2000; Weigel et al. 2000; Duehr et al. 2006; Fischer et al. 2010), demonstrating improvement relative to unbuffered reaches after as few as 3 years (Duehr et al. 2006). By contrast, however, new herbaceous buffers along agricultural headwater streams in Ohio were found to have only minimal impact on those stream ecosystems in the first 4 to 6 years after planting (Smiley et al. 2011). We expect that buffers established at SFWR sites are helping to reduce the effects of past and present-day agriculture on the stream ecosystem, and that these changes are already occurring in the short-term. However, additional forces may be at play to confound interpretation of our habitat data.

After buffer establishment, we expected stream channels to narrow and deepen as riparian vegetation became established and stabilized the previously eroding banks. Those outcomes are typical when stream riparia are protected (e.g., from livestock) and near-bank vegetation is allowed to recover (Hunter 1991; Lyons et al. 2000; Weigel et al. 2000; Fischer et al 2010). However, after buffers were established at all SFWR sites, stream channels widened significantly (19% increase), water depths remained unchanged, vegetation coverage on stream banks declined significantly as the area of bare bank doubled, and the length of vegetation overhanging the channel was reduced by one-half. Unlike the reductions observed in fine substrates and embeddedness, these all run contrary to typical instream and riparian changes observed after river and stream buffers are developed and/or protected from disturbance (Sovell et al. 2000; Weigel et al. 2000; Fischer et al. 2010). However, these changes are consistent with continued channel disturbance, possibly caused by more frequent flood events within the region since our initial stream surveys in 2005 (Keillor 2010; Mundahl and Hunt 2011).

Evidence is accumulating in support of the hypothesis that the magnitude and frequency of extreme precipitation and flooding events are increasing in the upper midwestern United States (Olsen et al. 1999; Andresen et al. 2012; Mallakpour and Villarini 2015). Early reports (Karl et al. 1995; Angel and Huff 1997) indicated that the frequency of heavy (> 5 cm/day) rainfall events in this region had increased by as much as 20% over the past century. More recent data for the region follow the same trends, demonstrating both a sharp increase in extreme (1-in-5 years) precipitation events after the year 2000 (Andresen et al. 2012) and providing strong evidence for increasing frequencies (but not greater magnitudes) for both heavy rainfall and flooding (Mallakpour and Villarini 2015). Many of the changes in instream and riparian conditions observed at SFWR sites between 2005 and 2018 could have resulted from more frequent, high streamflows which can scour the channel and banks, removing fine substrates while widening channels and reducing vegetation coverage on stream banks. These outcomes often are typical in stream and river systems exposed to frequent flooding (Waters 1995; Rosgen 2006; Allan et al. 2021). However, SFWR sites also demonstrated increased habitat (e.g., reduced dominance by runs, increased riffle abundance) and substrate diversity and nearly a doubling of instream vegetative cover for fish during 2018 surveys, opposite of how stream habitats normally respond to increased flooding and disturbance (Lorang and Hauer 2006; Rosgen 2006; Duehr et al. 2008). Increased baseflows of streams in this region (Ayers et al. 2018, 2021a, b) also could be responsible for the greater depths and stream widths and faster velocities observed at SFWR sites in 2018. Consequently, we cannot determine whether the habitat changes observed at SFWR sites were the result of buffer improvements, or if they resulted from frequent disturbance from flooding, greater base discharges, or some combination of these. Additional habitat surveys should be conducted in future years to determine the effectiveness of buffers in modifying stream habitats at the SFWR sites.

None of the benthic macroinvertebrate measures examined in this study of the SFWR displayed a significant change after buffer law implementation, with benthic communities rated as “poor” both before and after buffers were in place. Although some studies have shown that benthic macroinvertebrate assemblages respond positively to the presence (or increasing widths) of riparian buffers (Smiley et al. 2011; Albertson et al. 2018), others have reported little, if any, response of instream invertebrates to improved buffer conditions (Sovell et al. 2000; Weigel et al. 2000; Duehr et al. 2006). The lack of invertebrate response to riparian buffers has been variously attributed to insufficient time for biota to respond to improving instream conditions (Sovell et al. 2000), a reduced influence of invertebrates to riparian versus upstream watershed conditions (Weigel et al. 2000), or the simple lack of a source of potential invertebrate colonizers in watersheds with a long history of widespread agricultural disturbance (Fischer et al 2010; Mundahl and Mundahl 2022). Any or all of these may have played roles in our study of SFWR macroinvertebrates.

A time span of only 1 to 3 years after buffers were “finalized” at SFWR sites may simply have been too short for macroinvertebrates to exhibit a response to those buffers. Although stream water chemistry and physical habitats may respond quickly to development of riparian buffers (Sovell et al. 2000), longer time periods (e.g., 4 to 10 years) may be required after buffer installations before biotic assemblages exhibit observable changes related to riparian conditions (Weigel et al. 2000; Smiley et al. 2011). This delayed response often is attributed to the time required for organisms to disperse from nearby refugia or connected waters into the newly buffered reaches with improving physical conditions (Sovell et al. 2000; Smiley et al. 2011). However, benthic macroinvertebrates are capable of rapidly colonizing quality stream habitats in 2 years or less via downstream drift and aerial dispersal of emerging adult insects, especially for species capable of producing several generations per year (Waters 2000; Petersen et al. 2004; Mundahl and Hunt 2011; Trekels et al. 2011).

Benthic invertebrates also may not respond to localized buffers or other conservation efforts within agricultural drainages if upstream reaches remain unbuffered and unprotected, allowing watershed land uses to impact those stream segments (Nerbonne and Vondracek 2001). Consequently, the benefits of riparian buffers to macroinvertebrates may not be realized directly at the site of the buffer, but 100 to 300 m downstream from the protected reach (Weigel et al. 2000). However, these findings cannot be used to explain why SFWR invertebrate assemblages failed to respond to buffer development. Minnesota’s buffer rule mandates continuous buffer protection along the entire stream, eliminating the previous mosaic distribution of buffers that failed to protect streams from nearby agricultural activities (Sovell et al. 2000; Nerbonne and Vondracek 2001).

The lack of benthic macroinvertebrate response to new buffer placements in the SFWR could be caused simply by a failure of colonizing invertebrates to repopulate the stream sites from other localities. Our data suggest that the pool of possible macroinvertebrate colonizers within the SFWR is limited (consistently low taxa richness at all sites). Additionally, the historical impact of agriculture on this region has been severe (Waters 1995; Thorn et al. 1997; Trimble 2013), and the dispersal opportunities for invertebrate colonizers are limited due to topographical barriers (tall, wooded bluffs and distances > 15 km) that separate the SFWR from other less-impacted and more taxa-rich streams and rivers within the region (Mundahl and Mundahl 2022). Reintroducing additional invertebrate taxa via translocation from less-impacted systems (Grégoire Taillefer and Wheeler 2013) may be needed to facilitate the colonization process. Otherwise, macroinvertebrate diversity in the SFWR may remain limited and not indicative of stream habitat quality for several decades (Harding et al. 1998).

Fish assemblages showed no change after buffers were mandated at SFWR sites, even as some of the physical variables in their stream habitats were changing. Stream sites bordered by mature, well-developed buffers have been reported as having the highest fish IBI scores, the best fish community diversity, and the greatest fish taxa richness compared to unbuffered or more recently buffered locations (Duehr et al. 2006). But other studies (Sovell et al. 2000; Fischer et al. 2010; Smiley et al. 2011) have reported that fish communities responded only weakly, if at all, to stream buffer improvements. As with invertebrate communities, fish may require longer time periods (up to 10 years) encompassing multiple generations before their assemblages may reflect improved, present-day stream conditions (Sovell et al. 2000). River and stream fish communities are highly responsive to a wide range of environmental stressors and have long been used to evaluate the integrity of lotic systems (Karr 1981; Simon 1999, 2003). Coldwater fish communities in particular are sensitive to streamside conservation practices (Mundahl and Simon 1999; Wang et al. 2006), so the lack of change in fish communities after buffer establishment in the SFWR was unexpected.

Brown trout occur throughout the entire length of the SFWR, although only the lower 20.6 km (36%) of the system is managed as a trout stream (Minnesota Department of Natural Resources 2021). We anticipated that populations of brown trout would expand rapidly in stream reaches with improving habitats. Low-density brown trout populations can increase rapidly via high reproduction rates (Schnaser and Mundahl 2022), outcompeting other species and preying heavily on small forage fishes when they are available (Budy and Gaeta 2018). The condition (based on relative weight) of brown trout in the upper reaches of the SFWR was high, suggesting limited competition and abundant food resources, especially small fishes (Varela 2020). The low numbers of trout present and the lack of juvenile trout at many sites after buffer installation may indicate that stream habitats are adequate for trout growth and survival, but not yet suitable for spawning and/or egg incubation and embryo development. Brown trout require clean, coarse substrates for successful reproduction (Raleigh et al. 1986; Hunter 1991; Waters 1995), and many SFWR sites simply may lack sufficient suitable substrate to support either spawning or egg incubation, or both. Occasional brown trout spawning surveys conducted in the lower reaches of the SFWR (near sites 12 and 13) report spawning conditions sufficient to support strong (300 to 900 fish/km, 70 to 190 kg/hectare), self-sustaining brown trout populations (Anderson 1983; D. Dieterman, MN DNR, personal communication). Similar spawning surveys are needed to assess the ability of most of our SFWR sites to support trout reproduction, especially locations within the upstream one-third of the system.

In conclusion, implementation of a mandatory buffer law in Minnesota has produced promising improvements in some stream variables in the SFWR, relative to pre-buffer rule conditions. Coarser bottom substrates with reduced embeddedness and improved habitat heterogeneity indicate that stream sites are changing 1 to 3 years after riparian buffers were implemented. However, widening stream channels and reduced vegetative cover on stream banks suggest that continuing disturbances may be impacting stream habitat recovery after buffer implementation. Neither fish nor benthic macroinvertebrate communities have yet responded favorably to buffer improvements. Additional investigations at SFWR sites should focus on possible, ongoing improvements in stream habitat, as well as on the status of biotic communities, after buffers have “matured” to produce maximal protection for the stream system.

Acknowledgements We thank the private landowners who granted us permission to access the river for this study. T. Biederman, O. Graziano, N. Zwonitzer, and K. Nelson-Thomas assisted with field collections, and O. Graziano and N. Zwonitzer assisted with benthic invertebrate sample sorting and identification.

Author contributions NM and WV developed the concept for the paper. All authors participated in field work and benthic invertebrate sample sorting and identification. NM performed the analyses and led the writing of the paper. All authors have edited the manuscript and approved the submitted version.

Funding This project was supported by grants from the Minnesota Pollution Control Agency and Minnesota’s Environment and Natural Resources Trust Fund, as recommended by the Legislative-Citizen Commission on Minnesota Resources (M.L. 2017, Chp. 96, Sec. 2, Subd. 04d / Water Quality Monitoring in Southeastern Minnesota Trout Streams).

Compliance with ethical standards

Conflict of interest The authors declare no competing interests.

Ethics approval Fish surveys were made under special permits (Numbers 12769, 23668, 28374) from the Minnesota Department of Natural Resources, Division of Fish and Wildlife, Section of Fisheries, and were conducted with the approval of the Winona State University Institutional Animal Care and Use Committee (1317-64-1, 1310072-2).

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Stream Habitats and Aquatic Communities in an Agricultural Watershed: Responses to a Mandatory Riparian Buffer Law

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