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\begin{document}

\title{Mitigation of impacts of cattle access on stream ecosystems -- efficacy
of fencing}



\author[1]{M. O’Sullivan}%
\author[2]{Daire Ó hUallacháin}%
\author[3]{Patrícia  Antunes}%
\author[3]{Eleanor Jennings}%
\author[3]{Suzanne Linnane}%
\author[3]{Caroline Wynne}%
\author[1]{M. Kelly-Quinn}%
\affil[1]{University College Dublin School of Biology and Environmental Science}%
\affil[2]{Environment Research Centre Johnstown Castle}%
\affil[3]{Dundalk Institute of Technology Centre for Freshwater and Environmental Studies}%


\vspace{-1em}



  
  \date{May 21, 2024}


\begingroup
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\selectlanguage{english}
\begin{abstract}
Headwater streams can constitute up to 80\% of river channel length and
are vulnerable to anthropogenic pressures due to their high connectivity
to adjacent land, large relative catchment size and low dilution
capacity. In these environments unrestricted cattle access is a
potential significant cause of water quality deterioration, resulting
from increases in stream bank erosion, riparian damage and sediment
deposition among others. Several studies have reported improvements in
physico-chemical and hydromorphological conditions of streams following
elimination of cattle access; few, however, have focussed on the
ecological impacts of such management practices. Here, such impacts are
assessed. We look at the short-term effects by comparing habitat
condition, sediment deposition, and instream macroinvertebrate
communities upstream and downstream of cattle access points prior to,
and one year following exclusion via fencing. The long-term effects are
also measured by reassessing a small stream catchment entirely fenced
off from cattle access in 2008 under a concerted management effort. In
the short term, cattle exclusion led to reduction in deposited sediment
downstream of cattle access points and a related homogenisation of
macroinvertebrate community structure between upstream and downstream
sampling points. Increased abundances of specific indicator taxa (
\emph{Ancylus fluviatilis}, Glossosomatidae and Elmidae) in the fenced
catchment following 9 years of exclusion highlight the long-term
ecological benefits of such mitigation practices. These findings
highlight the importance of incentivised agri-environment schemes in
reducing the negative impacts of cattle access to these vulnerable
ecosystems.%
\end{abstract}%



\sloppy


\section*{Mitigation of impacts of cattle access on stream ecosystems --
efficacy of
fencing}

{\label{mitigation-of-impacts-of-cattle-access-on-stream-ecosystems-efficacy-of-fencing}}

O'Sullivan, M.\textsuperscript{1}, \selectlanguage{ngerman}Ó hUallacháin, D.\textsuperscript{3},
Antunes, P. O.\textsuperscript{4}., Jennings, E.\textsuperscript{4} ,
Suzanne Linnane\textsuperscript{4} , Caroline Wynne\textsuperscript{4} ,
\& Kelly-Quinn, M., \textsuperscript{1,2},

\textsuperscript{1} School of Biology and Environmental Science,
Freshwater Ecology Research Group, University College Dublin, Dublin,
Ireland, and \textsuperscript{2}UCD Earth Institute

\textsuperscript{3} Teagasc, Johnstown Castle Research Centre, Wexford,
Ireland

\textsuperscript{4} Centre for Freshwater and Environmental Studies,
Dundalk Institute of Technology, Dublin Road, Dundalk

Corresponding Author: Mary Kelly-Quinn -- mary.kelly-quinn@ucd.ie

\par\null

ABSTRACT

Headwater streams can constitute up to 80\% of river channel length and
are vulnerable to anthropogenic pressures due to their high connectivity
to adjacent land, large relative catchment size and low dilution
capacity. In these environments unrestricted cattle access is a
potential significant cause of water quality deterioration, resulting
from increases in stream bank erosion, riparian damage and sediment
deposition among others. Several studies have reported improvements in
physico-chemical and hydromorphological conditions of streams following
elimination of cattle access; few, however, have focussed on the
ecological impacts of such management practices. Here, such impacts are
assessed. We look at the short-term effects by comparing habitat
condition, sediment deposition, and instream macroinvertebrate
communities upstream and downstream of cattle access points prior to,
and one year following exclusion via fencing. The long-term effects are
also measured by reassessing a small stream catchment entirely fenced
off from cattle access in 2008 under a concerted management effort. In
the short term, cattle exclusion led to reduction in deposited sediment
downstream of cattle access points and a related homogenisation of
macroinvertebrate community structure between upstream and downstream
sampling points. Increased abundances of specific indicator taxa
(\emph{Ancylus fluviatilis} , Glossosomatidae and Elmidae) in the fenced
catchment following 9 years of exclusion highlight the long-term
ecological benefits of such mitigation practices. These findings
highlight the importance of incentivised agri-environment schemes in
reducing the negative impacts of cattle access to these vulnerable
ecosystems.

\textbf{Key Words:}

Cattle Access; Freshwater; Fencing; Macroinvertebrates

\subsection*{Introduction}

{\label{introduction}}

Agriculture, particularly in intensively managed systems, can pose a
risk to water quality via both diffuse and point sources. Unrestricted
cattle access to streams and rivers is among these sources and can have
potentially negative impacts on stream water quality as has been
demonstrated both in the temperate European setting (Conroy et al.,
2016a; O'Sullivan et al., 2019a \& b) and internationally (Braccia and
Voshell, 2007; Vidon et al., 2008; O'Callaghan et al. 2019, Raymond and
Vondracek, 2011). Facilitating animal access to watercourses provides
farmers with a cheap, low-maintenance source of water for their
livestock.

Cattle access to streams has been linked with increases in levels of
deposited sediment and habitat homogenisation (Herbst et al., 2012),
hydromorphological changes (Belsky et al., 1999), degraded riparian
conditions (Grudzinski et al., 2016), physico-chemical degradation
(Arnaiz et al., 2011), alterations in biological communities such as
aquatic macroinvertebrates (Braccia and Voshell, 2007; Conroy et al.
2016a) and increases in sediment concentrations of
\emph{Escherichiacoli} (Bragina et al., 2017).

Cattle access mediated changes in macroinvertebrate communities are
driven primarily by changes in stream hydromorphology and riparian
habitat conditions, and in elevated nutrient and sediment levels in
streams (Li et al., 1994; Stone et al., 2005; Arnaiz et al., 2011;
Miserendino et al., 2011; O'Callaghan et al. 2019). Increases in
limiting nutrients such as phosphorus and nitrogen, as a result of
cattle defecation both directly in streams and in the adjacent riparian
zone, can lead to concentrations that are detrimental in themselves
(Sarriquet et al., 2006), but can also result (in association with
reductions in riparian cover/shade) in proliferations of instream algal
communities (Braccia and Voshell, 2007). Damage to stream banks by the
erosive action of cattle hooves as they enter and exit streams, and as
they consume and trample streamside vegetation, contributes
significantly to stream suspended sediment inputs and bed loads of fine
sediment (Scrimgeour and Kendall, 2003; Braccia and Voshell, 2007;
Ranganath et al., 2009; Schulte et al., 2009; O'Sullivan et al.,
2019a,b; Rice et al., 2021). Habitat change as a result of algal
proliferation and sediment induced stream bed homogenisation seriously
alters macroinvertebrate communities in affected streams (Stone et al.,
2005; Sarriquet et al., 2006; Braccia and Voshell, 2007). The
environmental impact of cattle access to watercourses have been
recognised in agri-environmental policy, with cattle exclusion being
included in various schemes under Pillar I and Pillar II of the Common
Agricultural Policy (see Kilgarriff et al., 2020). Fenced riparian
buffer measures have been included in most European agri-environment
schemes (Dworak et al., 2009), including Ireland's, and are amongst the
commonest mitigation measures to prevent cattle access to watercourses.

Various methods of limiting livestock access to waterbodies have been
used and empirically tested. Miner et al. (1992), Bagshaw et al. (2008)
and Rawluk et al. (2014) studied the use of alternative water sources to
encourage cattle congregation away from streams and riparian areas.
Legrand et al. (2011) examined the use of ``cow showers'' as a means of
attracting cattle away from shaded riparian areas in the arid
Californian climate. The impacts of varying grazing practices have been
tested by Raymond and Vondracek (2011), where the benefits of reduced
grazing intensity on stream ecosystems via rotational grazing methods
were demonstrated. More commonly, however, has been the complete
exclusion of cattle from watercourses via fencing (Li et al., 1994;
Bewsell et al., 2007; Collins et al., 2010).

Various studies have cited how restriction of cattle access to streams
has positively affected factors such as instream and riparian habitat
(Weigel et al., 2000); stream hydromorphology (Agouridis et al., 2005),
reduced sediment loads (Collins et al., 2010) and physico-chemical
conditions (Batchelor et al., 2015; Line, 2003). Fewer studies have
examined the response of stream biota to restricted cattle access and
most of these studies are from North America (e.g. Raymond and
Vondracek, 2011 (macroinvertebrates), Derlet et al., 2017 (microbial)).
The efficacy of cattle exclusion as a means of improving stream water
ecological quality has not been widely assessed in the context of the
temperate northwest of Europe where pastoral farming predominates.

This study aimed to assess the environmental benefits of cattle
exclusion via fencing in the short-term (one-year post-fencing), by
examining changes in levels of deposited sediment, habitat condition and
macroinvertebrate communities in streams in Ireland. Additionally, a
case study on the longer-term responses to fencing (9 years) is also
presented. Here, 2008 macroinvertebrate data from two sub-catchments
(one which was entirely fenced as a part of a group water scheme in the
following year), are compared to data collected nine years following
exclusion. The results will help inform further development of
programmes of measures to improve water quality in agricultural
catchments.

\section*{2. Methods}

{\label{methods}}

\subsubsection*{2.1 Sites}

{\label{sites}}

Eight cattle access points located in five catchments in the east and
south of Ireland were sampled for macroinvertebrates and deposited
sediment prior to, and one year following exclusion of cattle from
streams via fencing (as a part of the Green Low Carbon Agri-environment
Scheme (GLAS)) in October 2016 and October 2017, respectively (Figure
1). Three sites (GLAS Sites 1, 2 and 3) were located along the moderate
status, second order Blacklion stream in Co. Carlow. Two sites (GLAS
Sites 4 and 5) were located on separate, good-high status tributaries of
the Munster Blackwater in Co. Cork. GLAS Site 6 was located on the poor
status Commons stream, a tributary of the White river in Co. Louth. GLAS
Site 7 was located on a good status, first order tributary of the Big
river in Co. Louth. GLAS Site 8 was on a moderate status tributary of
the Milltown Lake catchment in Co. Monaghan. Geology was variable (see
supplementary information) and across sites while catchment land use was
principally grassland. Stream substrates at GLAS Sites 1, 2, 4 and 7
were predominately gravel whereas coarse substrates, mainly cobble,
dominated at all other sites. Stream widths ranged from 0.72m to 3.60 m
(supplementary information).

The study on the long-term efficacy of fencing was carried out in the
Milltown Lake catchment of Co. Monaghan (Figure 1). Here 13 sites were
sampled in October 2008 (see Wynne and Linnane, 2008). Seven sites were
located in a sub-catchment fed from the Carnagh lake in Tievenamara
(TV), that was entirely fenced off from cattle following sampling in
2008, as part of remediation works for a group water scheme. Six sites
were located in a control catchment in the Tullycaghney (TH) that was
not fenced as a part of a concerted effort (although some ad-hoc,
localised fencing has taken place since sampling in 2008). All sites
were re-sampled in October 2017, nine years post fencing.

\subsubsection*{2.2 Macroinvertebrate
sampling}

{\label{macroinvertebrate-sampling}}

For the short-term fencing study, twelve macroinvertebrate samples were
collected at each site prior to fencing (in October 2016) and one-year
post fencing (in October 2017) using a standard
0.9m\textsuperscript{2}Surber sampler. Sampling points were located in
riffle geomorphic units immediately upstream (x 6) and downstream (x 6)
of cattle access points, hereafter referred to as \emph{control} and
\emph{pressure} points, respectively. A total of 96 samples (six Surbers
per sampling point) were collected in each year. The majority of insect
taxa were identified to species level; Coleoptera and Diptera were
identified to family level and oligochaetes were left at order level.

For the long-term fencing study, samples were collected (in 2008 and
2019) by 2-minute kick sampling from riffle habitats at each sampling
point. To enable comparisons with the pre-fencing study (Wynne and
Linnane, 2008) the sampling methods of Wynne and Linnane were replicated
as closely as possible and the macroinvertebrates were identified to
family level.

\subsubsection*{2.3 Sediment sampling}

{\label{sediment-sampling}}

Deposited sediment sampling was undertaken as part of the short-term
study only. The samples were taken from riffle habitats at control and
pressure points at the same time as macroinvertebrate sampling, using
the `Quorer' resuspendable sediment sampling technique (Lambert and
Walling, 1988; Conroy et al., 2016b). Six samples were collected at each
sampling point resulting in a total of 96 deposited sediment samples
from each sampling period.

The deposited sediment samples were filtered using pre-ashed Whatman
glass Microfibre GF/C filters following standard methods for suspended
sediment as set out by the American Public Health Association (APHA,
1995). They dried at 103\selectlanguage{english}0C to a constant weight and weighed to give
sediment mass (g). Organic matter mass was determined by weights of
filter papers upon loss on ignition of samples dried at 5500\selectlanguage{english}C to a
constant weight (APHA, 1995). Estimated resuspended sediment was
calculated as mg l\textsuperscript{-1} and converted to g
m\textsuperscript{-2} using the water volume within the stilling well
(Conroy et al., 2016b). Percent organic matter (\%OM) was calculated as
a percentage of overall sediment mass.

\subsubsection*{2.4 Habitat assessment}

{\label{habitat-assessment}}

Habitat assessment was carried out as part of the short-term study prior
to fencing (2016) and following one year of exclusion (2017). It
consisted of both quantitative measures of stream bank parameters and an
assessment of qualitative parameters, carried out at the reach scale. A
total of 13 qualitative sub-indices were calculated and were summed to
produce a Total Habitat Index (THI) and a Riparian Habitat Index (RHI)
(see O'Sullivan et al., 2019a). Each habitat assessment sub-index was
scored on an ordinal scale from 0-20 or 0-10 where stream banks were
assessed individually, with higher scores equating to better habitat
condition.

\subsubsection*{2.5 Data analysis}

{\label{data-analysis}}

Short-term study

A before, after, control and impact (BACI) study design was used to test
whether short-term exclusion of cattle via fencing affected levels of
deposited stream sediment metrics and macroinvertebrate community
structure. The analyses of macroinvertebrate data tested the hypothesis
that differences in community structure between control and pressure
points prior to fencing would be greater than those differences
following fencing and one year of cattle exclusion. This hypothesis was
tested in Primer using the PERMDISP function (Anderson, 2006).

Here, \emph{Site} (levels: 1-8) and \emph{Fencing} (levels: Pre and
Post) factors were combined to create the grouping factor for the
PERMDISP procedure. Thus, samples from control and pressure points were
combined for each site and the homogeneity of dispersions between
samples (taking in both control and pressure samples) prior to and
following fencing were tested via permutations, as is the protocol with
the PERMDISP procedure. The hypothesis, thus states that homogeneity of
dispersions will increase following fencing, or conversely, the within
group variation will be reduced following fencing (Figure 2).

Pairwise comparisons of PERMDISP results and mean \emph{z} scores were
assessed to identify any convergence or divergence of samples between
pre-fencing and post-fencing time periods at each site.

Potential differences in the multivariate locations (centroids) of
macroinvertebrate samples (community structure) based on control v
pressure differences before and after fencing were tested using
PERMANOVA. In these analyses \emph{Fencing} and \emph{Treatment} were
combined to create a fixed factor with four levels (Pre-fencing control,
pre-fencing pressure, post-fencing control and post-fencing
pressure).\emph{Site} was a random factor with eight levels (GLAS Sites
1, 2, 3, 4, 5, 6, 7 and 8).

SIMPER analysis was used to identify taxa that contributed to
dissimilarities in community structure between four combinations
of\emph{Fencing/Treatment} levels:

\begin{enumerate}
\tightlist
\item
  pre-fencing and post-fencing control points
\item
  pre-fencing and post-fencing pressure points
\item
  pre-fencing control and pre-fencing pressure points
\item
  post-fencing control and post-fencing pressure points.
\end{enumerate}

These SIMPER comparisons were made using data from across all eight
sites to identify taxa that demonstrated consistent changes to
exclusion.

Univariate metrics (Ephemeroptera Plecoptera Trichoptera (EPT)
abundance, \% EPT, richness indices, functional feeding group (FFG)
metrics and fine sediment sensitivity rating (FSSR) metrics) were also
analysed using this study design. Pairwise tests for pairs of
the\emph{Fencing/Treatment} term within levels of the \emph{Site} term
were carried out to investigate any interaction effects in relation to
univariate metrics.

Mean deposited sediment values were calculated for each Treatment level
(control and pressure) at each site for the pre-fencing and post-fencing
periods. PERMANOVA analysis was carried out using a two-factor design
with Fencing and Treatment as two fixed factors, each with two levels
(pre-fencing and post-fencing, and control and pressure, respectively).

With respect to habitat, qualitative habitat assessment scores were
analysed as univariate data using a two factor PERMANOVA (based on
Euclidean distance) with \emph{Fencing} and \emph{Treatment} as fixed
factors. Habitat scores were square root transformed prior to analysis.

\subsubsection*{2.6 Long-term study}

{\label{long-term-study}}

A blocked study design was used to assess for effects of long-term
fencing on macroinvertebrate community structure, using a three factor
PERMANOVA. The three factors were \emph{Fencing} and
\emph{Catchment}(both fixed with two levels; pre-fencing and
post-fencing and fenced (TV) and control (TH) respectively) and
\emph{Site} nested in\emph{Catchment} .

Blocked study designs as carried out in Primer v.6 are akin to repeated
measure PERMANOVAs and proceed with the exclusion of the `highest level'
interaction term (\emph{Site(Catchment)} x \emph{Fencing} in this case)
from the model (Anderson et al., 2008). Blocking of un-replicated
samples leads to greater power of detection (Fisher, 1935).

\section*{3. Results}

{\label{results}}

\subsection*{3.1 Short-term study}

{\label{short-term-study}}

\subsubsection*{3.1.1 Sediment mass and organic matter content: GLAS
sites}

{\label{sediment-mass-and-organic-matter-content-glas-sites}}

Mean deposited sediment levels in the pre-fencing period ranged from
19.96\selectlanguage{ngerman}± 5.29 gm\textsuperscript{-2} (SE) at the GLAS site 11 control
point to 538.15± 473.74 gm\textsuperscript{-2} (SE) at the GLAS site 3
pressure point. In the post fencing period, mean deposited sediment
levels ranged from 40.48± 11.94 gm\textsuperscript{-2} (SE) at the GLAS
site 7 control point to 458.76± 197.88 gm\textsuperscript{-2} (SE) at
the GLAS site 1 control point.

PERMDISP analysis detected a marginally significant difference in
homogeneity of dispersions among samples for levels of deposited
sediment mass (\emph{z} scores) at sampling points prior to fencing
compared to sampling points following fencing (F\textsubscript{15,188}=
P=0.06*), but none in relation to \%OM. No reductions were found,
however a pattern of reductions in dispersions (z scores) following
fencing was observed at several sites (Figure 3).

A significant interaction effect was identified between \emph{Fencing}
and \emph{Treatment} in relation to deposited sediment mass
(F\textsubscript{1,28}= 5.17, P=0.03). Pairwise results demonstrated
significantly greater masses of deposited sediment at pressure points
compared to control points, prior to fencing
(t\textsubscript{1,14}=2.38, P=0.04) but no significant difference
following fencing. It should be noted however, that there was also a
significant increase in levels of deposited sediment at control points
in year 2 compared to year 1 (t\textsubscript{1,14}=3.09, P=0.01)
(Figure 4).

There were no significant effects in relation to \%OM data for either
PERMDISP or PERMANOVA analyses.

\subsubsection*{3.1.2 Macroinvertebrate}

{\label{macroinvertebrate}}

A significant difference was detected in the magnitude of multivariate
dispersions among samples (across control and pressure points) between
pre-fencing and post-fencing periods (F\textsubscript{15,176}=3.81,
P=0.01). Pairwise comparisons indicated that at Site 1
(t\textsubscript{1,10}=2.34, P=0.04) and Site 3
(t\textsubscript{1,10}=3.58, P=0.01) (both on the Blacklion stream) the
magnitude of multivariate dispersion between samples was significantly
less following fencing. Additionally, a marginally significant decrease
in sample dispersions was found in relation to Site 4
(t\textsubscript{1,10}=2.07, P=0.06) (Figure 5).

Principle coordinate ordination (PCO) plots demonstrate (in most cases)
closer grouping of macroinvertebrate data between control and pressure
points post fencing (Figure 6).

SIMPER analysis highlighted twelve taxa that consistently contributed to
dissimilarities between the \emph{Treatment/Fencing} levels (Table
1).\emph{A. fluviatilis, R. semicolorata, S. pallipes} and
\emph{Agapetus}sp. had higher abundance at control points than at
pressures points in the pre-sampling period (Pre-Pressure v Pre-Control)
and greater in the post-fencing period compared to the pre-fencing
period at both control (Pre-Control v Post-Control) and pressure
(Pre-Pressure v Post-Pressure) points. The grazing riffle beetles,
\emph{Elmis aenea} and \emph{Limnius volckmari} occurred in higher
abundance at pressure points following fencing. Three taxa (\emph{B.
rhodani/atlaniticus} , Simuliidae and Chironomidae) were less abundant
in the post-fencing period at both control and pressure points, and less
abundant at control points in both the pre-fencing and post-fencing
periods.

In terms of the univariate metrics, there was a significant interaction
effect (\emph{Fencing/Treatment} x \emph{Site} ) in relation to total
richness (F\textsubscript{21,157}=5.10, P=0.01) and EPT richness
(F\textsubscript{21,157}=5.96, P=0.01). Pairwise results highlighted
significantly higher values at the pre-fencing control point relative to
the pre-fencing pressure point at Site 3 (t\textsubscript{1,10}=5.17,
P=0.01), Site 6 (t\textsubscript{1,10}=2.63, P=0.04) and Site 7
(t\textsubscript{1,10}=4.15, P=0.01), while at Site 1 total richness
values were greater at the pressure point (t\textsubscript{1,10}=3.92,
P=0.01) (Figure 7). Following fencing none of these significant
differences in total richness persisted and there were significant
increases in total richness at the pressure points following fencing
compared to prior to fencing at Site 2 (t\textsubscript{1,10}=4.13,
P=0.01), Site 3 (t\textsubscript{1,10}=6.29, P=0.01) and Site 6
(t\textsubscript{1,10}=2.79, P=0.03)

In relation to EPT richness (Figure 8), there were higher values at the
control points relative to the pressure points at Site 3
(t\textsubscript{1,10}=6.76, P=0.01) and Site 7
(t\textsubscript{1,10}=3.80, P=0.01) in the pre-fencing period. At Site
1, EPT richness values were greater at the pressure point in the
pre-fencing period (t\textsubscript{1,10}=3.39, P=0.02). Here again,
none of these differences persisted following fencing and there were
also significant increases in EPT richness at pressure points following
fencing at Site 2 (t\textsubscript{1,10}=5.26, P=0.01) and Site 3
(t\textsubscript{1,10}=5.26, P=0.01).

\subsubsection*{3.1.3 Habitat assessment: short-term
study}

{\label{habitat-assessment-short-term-study}}

Univariate PERMANOVA analysis of habitat index scores (THI and RHI) did
not highlight any difference between pre- and post-fencing periods.
Multivariate analysis of habitat score sub-indices and stream substrate
cover, geomorphic unit representation, physico-chemical measurements and
stream dimensions, similarly did not detect any differences. However,
RHI scores (available in supplementary information) did show an increase
at pressure points following fencing, but RHI scores were typically
higher at control points.

In relation to the univariate analyses, the \emph{ground cover}sub-index
of the RHI and THI was the only habitat parameter that showed a
difference between pre- and post-fencing periods. A
significant\emph{Fencing} and \emph{Treatment} interaction was observed
(F\textsubscript{1,28}=6.44, P=0.02) and pairwise results highlighted a
significantly (t\textsubscript{1,14}=2.65, P = 0.05) higher metric value
at post-fencing pressure points (9.75 ±0.66 SE) compared to pre-fencing
pressure points (7±0.2.65 SE). Significant main term
\emph{Treatment}effects with no \emph{Fencing} interaction were detected
for longitudinal connectivity (F\textsubscript{1,28}=9.29, P=0.01),
canopy cover (F\textsubscript{1,28}=7.58, P=0.01) and shrub layer cover
(F\textsubscript{1,28}=8.82, P=0.01). For each of these RHI sub-indices,
mean values at control points were greater than for those at pressure
points.

\subsection*{3.2 Long term study}

{\label{long-term-study-1}}

There was a significant interaction between the factors
\emph{Fencing}and \emph{Catchment} in relation to community structure
data in the Milltown Lake catchment (F\textsubscript{1,11}=2.78, P =
0.02). Significant differences were detected in community structure in
both the fenced (t\textsubscript{1,6}=3.09, P=0.01) and the control
(t\textsubscript{1,5}=2.51, P = 0.01) catchments between pre- and
post-fencing periods. Pairwise results also showed that prior to fencing
there was a significant difference in community structure between the
`to be fenced' catchment and control catchment
(t\textsubscript{1,11}=2.31, P = 0.01) that did not persist following
the fencing period.

Ordination based on principle components (PCO) illustrated a clear
separation between pre- and post-fencing sites in the fenced catchment
along the first PCO axes which accounted for 34.8\% of total variation
in the community structure (Figure 9). There is also a separation
between pre-and post-fencing samples in the control catchment, although
this appears not to be as pronounced as in the fenced catchment.

SIMPER analysis on the fenced catchment data showed that increased
abundances of Simuliidae, \emph{Ancylus fluviatilis,} Glossomatidae,
Elmidae and Baetidae, and reduced abundance of Asellidae, in the post
fencing period accounted for 41\% of the dissimilarity in community
structure between the pre- and post-fencing period. In the control
catchment 48\% of the dissimilarity in community structure between the
two periods was due to increased abundances of Gammaridae and Simuliidae
in the post-fencing period, and reduced abundances of
Baetidae,\emph{Hydropsyche siltalai} and Elmidae.

In relation to univariate metrics, there was a significant interaction
effect (\emph{Fencing} x \emph{Catchment} ) for \% ephemeropteran
abundance (\%E) (F\textsubscript{1,11}=19.62, P=0.01) and EPT abundance
(F\textsubscript{1,11}=14.69, P=0.01). Pairwise results for both, showed
significant differences in values between pre- and post-fencing periods
in both the fenced and control catchments. In the fenced catchment
values for both metrics generally increased following fencing while in
the control catchment values for both generally decreased (Figure 10).

\par\null

\section*{4. Discussion}

{\label{discussion}}

Degradation of freshwater resources with nutrients and sediment as a
result of agricultural practices is a key global challenge (Novotny,
1999; Schulte et al., 2009; EEA 2018; EPA 2020). Direct cattle access to
streams also represents a point source of nutrients, sediment and
pathogens (Godwin and Miner, 1996; Belsky et al., 1999; Davies\selectlanguage{english}-Colley et
al., 2004; Conroy et al., 2016a; Bragina et al., 2017, Rice et al.,
2021). In this study the efficacy of fencing as a means of reducing
in-stream sediment deposition and improving stream ecological quality in
both the short and longer term has been demonstrated.

Cattle exclusion over the short-term impacted instream deposited
sediment and aquatic macroinvertebrate communities

Significantly greater mean masses of deposited sediment were observed at
pressure points (relative to control points) prior to fencing, however,
following fencing there was no difference between control and pressure
points, despite a general increase in background levels of deposited
sediment. Patch scale deposits of fine sediment (as assessed here)
reflect stream bank erosion in the immediate upstream vicinity of the
affected site (Larsen et al., 2009). Habitat assessment results here did
not highlight any improvement in bank stability following exclusion of
cattle access, although increased ground cover in riparian areas was
observed. Reductions in the extent of bare ground following cattle
exclusion via fencing were also observed in Alberta, Canada by (Miller
et al., 2010a) and in the north of Ireland by Rice et al. (2021) and
were linked to improved water quality (Miller et al., 2010a). These
authors also found no improvement in stream bank stability following
four years of cattle exclusion but did suggest that increased ground
cover could also contribute to reduced rates of stream bank erosion.
Collins et al. (2010) concluded that stream bank erosion was
significantly reduced in all of six catchments studied in the UK
following fencing and eight years of exclusion, leading to reduced
siltation of salmonid spawning gravels. Hansen and Budy (2011) however
did observe significant improvement in stream bank stability after only
one to two years of exclusion of cattle via fencing in Utah. The
time-frame of the current study may have been too short to allow for
improvements in bank stability following fencing.

Elevated levels of deposited sediment have been shown to drive changes
in macroinvertebrate communities affected by cattle access (Braccia and
Voshell, 2006; Conroy et al. 2018). The results here have highlighted
the elimination of pre-fencing disparities in masses of deposited
sediment between control and pressure points as a result of one year of
exclusion via fencing. Additionally, reductions in multivariate
dispersions (macroinvertebrate data) between control and pressure
samples following fencing have been highlighted at several sites.
Warwick and Clarke (1993) found that multivariate dispersion among
marine community samples collected from impacted areas may be greater
than from un-impacted areas, thus, here, higher dispersion among
(control and pressure) samples prior to fencing compared to post-fencing
may indicate reduced community stress following restoration via fencing.
Other authors (see White et al., 2017) have shown increases in the
magnitude of multivariate dispersion following restoration of streams
and instream habitats due to improved habitat heterogeneity. It should
be noted however that multivariate dispersions in the current study were
measured across the gradient of impact (i.e. upstream and downstream of
cattle access points) and thus reductions in dispersion here represent a
convergence of communities towards \emph{control} conditions following
the removal of the cattle access impact. Results for EPT richness and
total richness also suggested that improvements in ecological quality
had occurred following one year of cattle exclusion at certain sites.

Cattle exclusion had a significant impact on grazing ephemeropteran,
plecopteran and trichopteran taxa

SIMPER results highlighted consistency in the taxa that contributed to
dissimilarities between the communities in the pre and post fencing
periods. The grazing EPT taxa, \emph{R. semicolorata, B. rhodani, Silo
pallipes} , \emph{Agapetus} sp. all contributed significantly to these
dissimilarities and these taxa generally showed consistent responses
(except \emph{B. rhodani} ). During the post-fencing sampling
season,\emph{R. semicolorata} , \emph{S. pallipes} and \emph{Agapetus}
sp. increased in abundances at both control and pressure points (across
all sites) compared to the pre-fencing period. Pre-fencing, these taxa
occurred in greater abundances at control than pressure points.
Interestingly however, this pattern was reversed in the post-fencing
period with abundances at pressure points increasing beyond those seen
at control points. Such a response may indicate that affected streams
are in a state of intermediate disturbance (Connell, 1978), with
community niches developing due to an abundance of food resources
(algae) for grazing taxa following a reduction of the main community
stressor, sediment. Cattle access points have been shown to negatively
affect grazers and EPT fauna such as \emph{Agapetus} sp. and \emph{R.
semicolorata} with sediment highlighted as the most likely stressor
(O'Sullivan unpublished). Similar impacts have also been highlighted by
Braccia and Voshell (2006) and Burdon et al. (2013). Elevated deposited
sediment can affect periphytic algal biomass and reduce its palatability
(Wood and Armitage, 1997; Schofield et al., 2004; Jones et al., 2012).
Higher fine sediment deposits downstream of cattle access points have
been detected in first and second order streams in the Irish setting
(O'Sullivan et al., 2019a) and were demonstrated here in the pre-fencing
sampling period. Such deposits however are transient in nature and can
be removed by high flows (Gomi et al., 2005). Nonetheless,
macroinvertebrate community composition reflects prevailing stream
conditions (Metcalfe, 1989) and sediment inputs from recurrent entry of
cattle to streams may represent a continuous stressor for these
communities.

Recovery time in macroinvertebrate communities from stressor effects is
variable and dependent on the specific stressor or stressors shaping
affected communities and the aspects of the communities assessed
(Laasonen et al., 1998; Muotka et al., 2002). Several authors have
highlighted similarly rapid responses in macroinvertebrate communities
to restoration efforts although not specifically relating to the
exclusion of cattle via fencing. Miller et al. (2010b) concluded from a
meta-analysis of a range of restoration projects that improved
macroinvertebrate richness can occur in communities within one year of
the commencement of restorative works. Contrastingly, Laasonen et al.
(1998) demonstrated a delayed recovery of shredder macroinvertebrates in
streams channelized for logging activities, relating the delay to poor
retention of coarse particulate organic matter (CPOM) in such streams.
In relation to cattle exclusion studies, recovery times are equally
variable with this variability reflecting, again, the main stressor
influencing communities and the scale of exclusionary measures. Herbst
et al. (2012) reported a relatively rapid recovery of macroinvertebrate
communities (in streams where bank erosion and sediment deposition were
significant drivers) following fencing (4 years), however with active
restoration (involving stream channel adjustment), even quicker
responses (2 years) were observed (Herbst and Kane, 2009). Where
significant depletion of riparian vegetation occurs however, greater
recovery times can be expected, owing to the substantial time required
for woody vegetation to re-establish and grow (Belsky et al., 1999;
Braccia and Voshell, 2007; Ranganath et al., 2009; Herbst et al., 2012).

Rapid responses of macroinvertebrates to restorative efforts are
indicative of communities that possess good elasticity (Halpern, 1988).
Greathouse et al. (2005) state that, recovery is driven by nearby
sources of macroinvertebrates for colonisation and the mobility of
dominant taxa. This point is reiterated by Diaz Villanueva et al. (2017)
who state that the distribution of invertebrates in disturbed
environments is shaped by the dispersal capabilities of the resident
taxa and the changes in environmental conditions along a spatial
gradient. Wallace (1990) cites the importance of drifting taxa and
aerially dispersing taxa (particularly in headwater streams) for
ecological recovery. Here eight out of the twelve taxa listed in Table 1
have an adult life stage that disperses aerially. Also, the downstream
extent of impacts of cattle access points in headwater streams is
limited (O'Sullivan et al., 2019a), thus the rapid restoration of
communities observed here is likely aided by both drifting taxa from
upstream control sites and from aerial dispersal of adults from
un-impacted downstream sites.

Fencing over the long-term results in improvements in macroinvertebrate
communities

In the present study, significant improvements in stream
macroinvertebrate community structure were seen in response to the
long-term fencing of a sub-catchment of the Milltown Lake in Co.
Monaghan. This was attributed to increases in the abundances of taxa
such as \emph{A. fluviatilis} and \emph{Glossomatidae} (taxa that
responded positively to fencing in the short term) and Elmidae (a taxon
that is shown to be sensitive to cattle access (Braccia and Voshell,
2007) and was more abundant downstream of access points following short
term fencing) in the post fencing sampling period of the fenced
catchment. These results are more emphatic given the apparent decrease
in ecological water quality in the control stream, where cattle access
was still permitted over the nine-year period. The lower ecological
quality scores in the post-fencing period were driven by increases in
taxa such as Simuliidae and Gammaridae and reductions in
Baetidae,\emph{Hydropsyche siltalai} and Elmidae compared to the
pre-fencing period. Despite Simuliidae and Gammaridae being highly
sensitive and moderately sensitive to sediment respectively, they are
also moderately tolerant of organic pollution and increases in their
abundances may reflect such pollution (Herbst et al., 2012; Burdon et
al., 2013). Furthermore, previous work by O'Sullivan et al. (2019a)
demonstrated that deposited sediment levels were unaffected by cattle
access at sites located in the control catchment, while a detailed
characterisation of the catchment by Wynne and Linnane (2008) concluded
that bare, coarse substrates were dominant in the catchment. Also work
by Bragina et al. (2017) in 2012-2013 found significantly higher
\emph{Escherichia coli}levels in the stream sediments for the unfenced
(control) tributary (TH) compared to the fenced tributary (TV),
suggesting higher loading with organic matter from animal and/or human
wastes.

The results from the univariate analyses also highlighted ecological
improvements in the fenced catchment and deteriorations in the control
catchment following the fencing period. The scope for inferences in
relation to the impact of fencing however, are limited due to the lack
of temporal replication over the course of the nine years between
fencing and the post-fencing sampling period.

Limitations in the study design are acknowledged by the authors,
particularly in relation to the failure of the analyses to account for
fluctuations in environmental conditions (e.g. sediment) and resultant
macroinvertebrate responses over time. Replication of sampling in
relation to the \emph{Fencing} factor (i.e. replicate sampling of
sediment and macroinvertebrate populations at multiple times prior to
fencing and after fencing, both at control and pressure points) would
provide a more robust basis on which conclusions could be drawn
(Stewart-Oaten et al., 1986; Miller et al., 2010b). Macroinvertebrate
populations, however, respond to the prevailing environmental conditions
and as such changes in macroinvertebrate metrics here are considered to
reflect changes in sediment deposition in the longer term.

Multiple targeted mitigation measures should be incorporated and
incentivised in catchment management plans

Results from this study support fencing as an effective method of
improving the ecological integrity of headwater streams. However, such
measures can be costly (Byers et al., 2005; Bayley and Li, 2008; Rawluk
et al., 2014), impractical (Sheffield et al., 1997) and unattractive to
farmers and other land managers (Graz et al., 2012). To avoid a
one-size-fits-all approach, Schulte et al. (2009) promoted targeting
fencing schemes to critical source areas, while Kilgarriff et al (2020)
found that fencing areas according to agricultural intensity to be a
cost-effective solution. Similarly, fencing and cattle exclusion alone
may not be sufficient to restore the ecological condition of impacted
watercourse and thus should not be considered a cure-all, but rather one
measure in a suite of mitigation measures that could be integrated into
wider management plans (Brannan et al., 2000; Schulte et al., 2009;
Collins et al., 2010; Smolders et al., 2015).Further research on the
socio-economic costs of fencing and costs associated with provision of
an alternative water supply (if necessary) as part of integrated
catchment management plans is required.

\par\null

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\section*{Figure Legends}

{\label{figure-legends}}

\textbf{Figure 1} Site locations for sites used to assess the short-term
(site names with the GLAS prefix) and long-term (TH (unfenced) and TV
(fenced) site name prefixes) efficacy of cattle exclusion.

\textbf{Figure 2} Experimental design demonstrating how the PERMDISP
routine was applied in this study. PERMDISP tests the hypothesis that
dispersion among triangles (pre-fencing) is greater than dispersion
among squares (post-fencing). The black lines (a) represent an
approximation of the dispersion/z-scores. The barbed wire represents
fencing (b).

\textbf{Figure 3} Measures of dispersions (z scores) between masses of
deposited sediment samples among control and pressure points during the
pre-fencing and post-fencing period.

\textbf{Figure 4} Mean deposited sediment levels (+- SE) at control and
pressure points prior to and following fencing across all sites. Letters
(A and B) identify significant \emph{Treatment} effects within a given
\emph{Fencing}period. Asterisks (* and **) identify significant
\emph{Fencing} effects within a given \emph{Treatment} .

\textbf{Figure 5} Measures of dispersions (z scores) among control and
pressure macroinvertebrate samples during the pre-fencing and
post-fencing period. Differing letters highlight where significant
differences occur

\textbf{Figure 6} PCO plots for macroinvertebrate community structure
data (based on Bray-Curtis dissimilarity matrices) for Site 1-8
(labelled a-h).PERMANOVA analysis of macroinvertebrate community
abundance (structure) highlighted a significant \emph{Fencing/Treatment}
x \emph{Site} interaction (F\textsubscript{21,157}=4.80, P=0.0002).

\textbf{Figure 7} Total richness values (+- SE) at control and pressure
points during the pre-fencing and post-fencing period Significant
differences between levels of the Treatment factor are designated by
differing letters (A and B), while differences between levels of the
Fencing factor are designated by differing numbers of asterisks' (* and
**).

\textbf{Figure 8} EPT richness values (+- SE) at control and pressure
points during the pre-fencing and post-fencing period. Significant
differences between levels of the Treatment factor are designated by
differing letters (A and B), while differences between levels of the
\emph{Fencing} factor are designated by differing numbers of asterisks'
(* and **).

\textbf{Figure 9} PCO plot of macroinvertebrate community structure data
showing greater separation of pre-fencing control and pressure samples
compared to post-fencing control and pressure samples.

\textbf{Figure 10.} Values of \%E abundance and EPT richness for samples
in the fenced (a and c) and control catchments (b and d).

\section*{Table}

{\label{table}}

\textbf{Table 1.} SIMPER results for selected taxa. Each column
indicates the relative change in abundance (indicated by directional
arrow) of the specified taxon at the second
listed\emph{Fencing/Treatment} level compared to the first. The
percentage contribution of each taxa to dissimilarities between the
given\emph{Fencing/Treatment} levels is given next to the directional
arrow. Control and Pressure refer to the sampling points.\selectlanguage{english}
\begin{longtable}[]{@{}llllllllll@{}}
\toprule
& Pre-Control v Post-Control & Pre-Control v Post-Control & Pre-Pressure
v Post-Pressure & Pre-Pressure v Post-Pressure & Pre-Pressure v
Pre-Control & Pre-Pressure v Pre-Control & Post-Pressure v Post-Control
& Post-Pressure v Post-Control & Cumulative \% contribution to
dissimilarities across Fencing/Treatment comparisons\tabularnewline
\emph{Ancylus fluviatilis} & \textbf{|} & \textbf{5.91\%} & \textbf{|} &
\textbf{7.71\%} & \textbf{|} & \textbf{3.02\%} & \textbf{|} &
\textbf{7.55\%} & 24.19\%\tabularnewline
\emph{Rhithrogena semicolorata} & \textbf{|} & \textbf{4.77\%} &
\textbf{|} & \textbf{4.97\%} & \textbf{|} & \textbf{3.60\%} & \textbf{|}
& \textbf{4.05\%} & 17.39\%\tabularnewline
\emph{Baetis rhodani} & \textbf{|} & \textbf{4.28\%} & \textbf{|} &
\textbf{4.90\%} & \textbf{|} & \textbf{5.56\%} & \textbf{|} &
\textbf{4.51\%} & 19.25\%\tabularnewline
\emph{Silo pallipes} & \textbf{|} & \textbf{3.74\%} & \textbf{|} &
\textbf{5.23\%} & \textbf{|} & \textbf{3.20\%} & \textbf{|} &
\textbf{5.10\%} & 17.27\%\tabularnewline
\emph{Agapetus} sp. & \textbf{|} & \textbf{4.04\%} & \textbf{|} &
\textbf{4.95\%} & \textbf{|} & \textbf{4.48\%} & \textbf{|} &
\textbf{5.33\%} & 18.80\%\tabularnewline
Hydropsychidae sp. & \textbf{|} & \textbf{3.89\%} & \textbf{|} &
\textbf{3.16\%} & \textbf{|} & \textbf{3.11\%} & \textbf{|} &
\textbf{3.95\%} & 14.11\%\tabularnewline
Simuliidae & \textbf{|} & \textbf{5.42\%} & \textbf{|} & \textbf{6.79\%}
& \textbf{|} & \textbf{7.22\%} & \textbf{|} & \textbf{5.86\%} &
25.29\%\tabularnewline
Chironomidae & \textbf{|} & \textbf{6.37\%} & \textbf{|} &
\textbf{6.20\%} & \textbf{|} & \textbf{8.24\%} & \textbf{|} &
\textbf{5.20\%} & 26.01\%\tabularnewline
Pediciidae & \textbf{|} & \textbf{3.17\%} & \textbf{|} & \textbf{3.36\%}
& \textbf{|} & \textbf{4.05\%} & \textbf{|} & \textbf{2.04\%} &
12.62\%\tabularnewline
\emph{Limnius volckmari} & \textbf{|} & \textbf{5.21\%} & \textbf{|} &
\textbf{4.03\%} & \textbf{|} & \textbf{4.80\%} & \textbf{--} &
\textbf{4.57\%} & 18.61\%\tabularnewline
\emph{Elmis aenea} & \textbf{|} & \textbf{4.23\%} & \textbf{|} &
\textbf{4.31\%} & \textbf{|} & \textbf{3.96\%} & \textbf{|} &
\textbf{4.59\%} & 17.09\%\tabularnewline
\emph{Gammarus duebeni} & \textbf{--} & \textbf{5.00\%} & \textbf{|} &
\textbf{4.61\%} & \textbf{|} & \textbf{6.00\%} & \textbf{|} &
\textbf{5.24\%} & 20.85\%\tabularnewline
Cumulative \% contribution to dissimilarities of all taxa for each
comparison & 56.03\% & 56.03\% & 60.22\% & 60.22\% & 57.24\% & 57.24\% &
57.99\% & 57.99\% &\tabularnewline
\bottomrule
\end{longtable}\selectlanguage{english}
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\textbf{Hosted file}

\verb`Table 1.docx`  available at \url{https://authorea.com/users/588330/articles/625577-mitigation-of-impacts-of-cattle-access-on-stream-ecosystems-efficacy-of-fencing}

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